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United States
Air Quality Criteria for
Ozone and Related
Photochemical Oxidants
(First External Review Draft)
Volume III of
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EPA/600/R-05/004CA
January 2005
Air Quality Criteria for Ozone and Related
Photochemical Oxidants
Volume
National Center for Environmental Assessment-RTF Office
Office of Research and Development
U.S. Environmental Protection Agency
Research Triangle Park, NC
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DISCLAIMER
This document is an external review draft for review purposes only and does not
constitute U.S. Environmental Protection Agency policy. Mention of trade names or
commercial products does not constitute endorsement or recommendation for use.
PREFACE
National Ambient Air Quality Standards (NAAQS) are promulgated by the United States
Environmental Protection Agency (EPA) to meet requirements set forth in Sections 108 and 109
of the U.S. Clean Air Act (CAA). Sections 108 and 109 require the EPA Administrator (1) to
list widespread air pollutants that reasonably may be expected to endanger public health or
welfare; (2) to issue air quality criteria for them that assess the latest available scientific
information on nature and effects of ambient exposure to them; (3) to set "primary" NAAQS to
protect human health with adequate margin of safety and to set "secondary" NAAQS to protect
against welfare effects (e.g., effects on vegetation, ecosystems, visibility, climate, manmade
materials, etc); and (5) to periodically review and revise, as appropriate, the criteria and NAAQS
for a given listed pollutant or class of pollutants.
In 1971, the U.S. Environmental Protection Agency (EPA) promulgated National Ambient
Air Quality Standards (NAAQS) to protect the public health and welfare from adverse effects of
photochemical oxidants. The EPA promulgates the NAAQS on the basis of scientific
information contained in air quality criteria issued under Section 108 of the Clean Air Act.
Following the review of criteria as contained in the EPA document, Air Quality Criteria for
Ozone and other Photochemical Oxidants published in 1978, the chemical designation of the
standards was changed from photochemical oxidants to ozone (O3) in 1979 and a 1-hour O3
NAAQS was set. The 1978 document focused primarily on the scientific air quality criteria for
O3 and, to a lesser extent, on those for other photochemical oxidants such as hydrogen peroxide
and the peroxyacyl nitrates, as have subsequent revised versions of the ozone document.
To meet Clean Air Act requirements noted above for periodic review of criteria and
NAAQS, the O3 criteria document, Air Quality Criteria for Ozone and Other Photochemical
Ill-ii
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Oxidants, was next revised and then released in August 1986; and a supplement, Summary of
Selected New Information on Effects of Ozone on Health and Vegetation, was issued in January
1992. These documents were the basis for a March 1993 decision by EPA that revision of the
existing 1-h NAAQS for O3 was not appropriate at that time. That decision, however, did not
take into account some of the newer scientific data that became available after completion of the
1986 criteria document. Such literature was assessed in the next periodic revision of the O3 air
quality criteria document, which was completed in 1996 and provided scientific bases supporting
the setting by EPA in 1997 of an 8-h O3 NAAQS that is currently in force together with the 1-h
O3 standard.
The purpose of this revised air quality criteria document for O3 and related photochemical
oxidants is to critically evaluate and assess the latest scientific information published since that
assessed in the above 1996 Ozone Air Quality Criteria Document (O3 AQCD), with the main
focus being on pertinent new information useful in evaluating health and environmental effects
data associated with ambient air O3 exposures. However, some other scientific data are also
presented and evaluated in order to provide a better understanding of the nature, sources,
distribution, measurement, and concentrations of O3 and related photochemical oxidants and
their precursors in the environment. The document assesses pertinent literature available
through 2004.
The present draft document (dated January 2005) is being released for public comment and
review by the Clean Air Scientific Advisory Committee (CASAC) to obtain comments on the
organization and structure of the document, the issues addressed, the approaches employed in
assessing and interpreting the newly available information on O3 exposures and effects, and the
key findings and conclusions arrived at as a consequence of this assessment. Public comments
and recommendations will be taken into account making any appropriate further revisions to this
document for incorporation into a Second External Review Draft. That draft will be released for
further public comment and CASAC review before last revisions are made in response and
incorporated into a final version to be completed by early 2006. Evaluations contained in the
present document will be drawn on to provide inputs to associated PM Staff Paper analyses
prepared by EPA's Office of Air Quality Planning and Standards (OAQPS) to pose options for
consideration by the EPA Administrator with regard to proposal and, ultimately, promulgation of
decisions on potential retention or revision, as appropriate, of the current O3 NAAQS.
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Preparation of this document was coordinated by staff of EPA's National Center for
Environmental Assessment in Research Triangle Park (NCEA-RTP). NCEA-RTP scientific
staff, together with experts from other EPA/ORD laboratories and academia, contributed to
writing of document chapters. Earlier drafts of document materials were reviewed by non-EPA
experts in peer consultation workshops held by EPA. The document describes the nature,
sources, distribution, measurement, and concentrations of O3 in outdoor (ambient) and indoor
environments. It also evaluates the latest data on human exposures to ambient O3 and
consequent health effects in exposed human populations, to support decision making regarding
the primary, health-related O3 NAAQS. The document also evaluates ambient O3 environmental
effects on vegetation and ecosystems, man-made materials, and surface level solar UV radiation
flux and global climate change, to support decision making on secondary O3 NAAQS.
NCEA acknowledges the valuable contributions provided by authors, contributors, and
reviewers and the diligence of its staff and contractors in the preparation of this draft document.
Ill-iv
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Air Quality Criteria for Ozone and Related
Photochemical Oxidants
(First External Review Draft)
VOLUME I
Executive Summary E-l
1. INTRODUCTION 1-1
2. PHYSICS AND CHEMISTRY OF OZONE IN THE ATMOSPHERE 2-1
CHAPTER 2 ANNEX (ATMOSPHERIC PHYSICS/CHEMISTRY) AX2-1
3. ENVIRONMENTAL CONCENTRATIONS, PATTERNS, AND
EXPOSURE ESTIMATES 3-1
CHAPTER 3 ANNEX (AIR QUALITY AND EXPOSURE) AX3-1
III-v
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Air Quality Criteria for Ozone and Related
Photochemical Oxidants
(First External Review Draft)
(cont'd)
VOLUME II
4. DOSIMETRY, SPECIES HOMOLOGY, SENSITIVITY, AND
ANIMAL-TO-HUMAN EXTRAPOLATION 4-1
CHAPTER 4 ANNEX (DOSIMETRY) AX4-1
5. TOXICOLOGICAL EFFECTS OF OZONE AND RELATED
PHOTOCHEMICAL OXIDANTS IN LABORATORY ANIMALS
AND IN VITRO TEST SYSTEMS 5-1
CHAPTER 5 ANNEX (ANIMAL TOXICOLOGY) AX5-1
6. CONTROLLED HUMAN EXPOSURE STUDIES OF OZONE AND
RELATED PHOTOCHEMICAL OXIDANTS 6-1
CHAPTER 6 ANNEX (CONTROLLED HUMAN EXPOSURE) AX6-1
7. EPIDEMIOLOGICAL STUDIES OF HUMAN HEALTH EFFECTS
ASSOCIATED WITH AMBIENT OZONE EXPOSURE 7-1
CHAPTER 7 ANNEX (EPIDEMIOLOGY) AX7-1
8. INTEGRATIVE SYNTHESIS 8-1
Ill-vi
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Air Quality Criteria for Ozone and Related
Photochemical Oxidants
(First External Review Draft)
(cont'd)
VOLUME
9. ENVIRONMENTAL EFFECTS: OZONE EFFECTS ON
VEGETATION AND ECOSYSTEMS 9-1
10. TROPOSPHERIC OZONE EFFECTS ON UV-B FLUX AND
CLIMATE CHANGE PROCESSES 10-1
11. EFFECT OF OZONE ON MAN-MADE MATERIALS 11-1
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Table of Contents
Page
List of Tables I-xiv
List of Figures I-xvi
Authors, Contributors, and Reviewers I-xviii
U.S. Environmental Protection Agency Project Team for Development of Air
Quality Criteria for Ozone and Related Photochemical Oxidants I-xix
U.S. Environmental Protection Agency Science Advisory Board (SAB)
Staff Office Clean Air Scientific Advisory Committee (CASAC)
Ozone Review Panel I-xx
Abbreviations and Acronyms I-xxi
9. ENVIRONMENTAL EFFECTS: OZONE EFFECTS ON VEGETATION
AND ECOSYSTEMS 9-1
9.1 INTRODUCTION 9-1
9.2 METHODOLOGIES USED IN VEGETATION RESEARCH 9-2
9.2.1 Introduction 9-2
9.2.2 Methods Involving Experimental Exposures to O3 9-3
9.2.2.1 "Indoor", Controlled Environment, and
Greenhouse Chambers 9-3
9.2.2.2 Field Chambers 9-5
9.2.2.3 Plume Systems 9-7
9.2.2.4 Comparative Studies 9-10
9.2.2.5 Ozone Generation Systems 9-11
9.2.2.6 Experimental Exposure Protocols 9-11
9.2.3 Methods Involving Exposures to O3 in Ambient Air 9-12
9.2.3.1 Air-exclusion Systems 9-13
9.2.3.2 Natural Gradients 9-13
9.2.3.3 Use of Chemical Protectants 9-14
9.2.3.4 Biomonitoring 9-16
9.2.3.5 Calibrated Passive Monitors 9-24
9.2.4 Numerical/Statistical Methodologies 9-25
9.2.5 Improved Methods for Defining Exposure 9-27
9.3 SPECIES RESPONSE/MODE-OF-ACTION 9-28
9.3.1 Introduction 9-28
9.3.2 Mechanisms of Ozone-Induced Plant Alterations 9-31
9.3.2.1 Changes in Metabolic Processes 9-32
9.3.2.2 Modifications of Plant Physiological Processes 9-34
9.3.3 Ozone Uptake by Leaves 9-35
9.3.3.1 Possible Reactions Within the Leaf 9-43
9.3.3.2 Toxicants Within the Wall Space 9-46
9.3.3.3 Products of Ozone 9-48
9.3.3.4 Antioxidants Within the Apoplastic Space 9-55
9.3.4 Wounding and Pathogen Attack 9-64
9.3.4.1 Peroxidases 9-67
9.3.4.2 Jasmonic Acid and Salicylic Acid 9-69
9.3.4.3 Stress-Induced Alterations in Gene Expression 9-69
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Table of Contents
(cont'd)
Page
9.3.5 Primary Assimilation by Photosynthesis 9-73
9.3.5.1 Photooxidation: Light Reactions 9-73
9.3.6 Alteration of Rubisco by Ozone: Dark Reactions 9-74
9.3.7 Carbohydrate Transformations and Allocation 9-77
9.3.7.1 Lipid Synthesis 9-79
9.3.8 Role of Age and Size Influencing Response to Ozone 9-82
9.3.9 Summary 9-84
9.4 MODIFICATION OF FUNCTIONAL AND GROWTH RESPONSES 9-85
9.4.1 Introduction 9-85
9.4.2 Genetics 9-88
9.4.2.1 Genetic Basis of O3 Sensitivity 9-90
9.4.3 Environmental Biological Factors 9-91
9.4.3.1 Oxidant-Plant-Insect Interactions 9-92
9.4.3.2 Oxidant-Plant-Pathogen Interactions 9-95
9.4.3.3 Oxidant-Plant-Symbiont Interactions 9-100
9.4.3.4 Oxidant-Plant-Plant Interactions: Competition 9-101
9.4.4 Physical Factors 9-104
9.4.4.1 Light 9-104
9.4.4.2 Temperature 9-108
9.4.4.3 Humidity and Surface Wetness 9-111
9.4.4.4 Drought and Salinity 9-113
9.4.5 Nutritional Factors 9-116
9.4.6 Interactions with Other Pollutants 9-118
9.4.6.1 Oxidant Mixtures 9-119
9.4.6.2 Sulfur Dioxide 9-119
9.4.6.3 Nitrogen Oxides, Nitric Acid Vapor, and Ammonia .... 9-120
9.4.6.4 Hydrogen Fluoride and Other Gaseous Pollutants 9-123
9.4.6.5 Acid Deposition 9-123
9.4.6.6 Heavy Metals 9-125
9.4.6.7 Mixtures of Ozone with Two or More Pollutants 9-125
9.4.7 Interactions with Agricultural Chemicals 9-125
9.4.8 Factors Associated with Global Climate Change 9-126
9.4.8.1 Ozone-Carbon Dioxide-Temperature Interactions 9-128
9.4.8.2 Ozone-UV-B Interactions 9-151
9.4.8.3 Interactions of Ozone with Multiple Climate
Change Factors 9-153
9.4.9 Summary - Environmental Factors 9-153
9.5 EFFECTS-BASED AIR QUALITY EXPOSURE- AND
DOSE-RESPONSE INDICES 9-157
9.5.1 Introduction 9-157
9.5.2 Summary of Conclusions from the Previous Criteria Document .... 9-158
9.5.3 Use of Exposure Indices to Establish Exposure-Response
Relationships 9-161
Ill-ix
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Table of Contents
(cont'd)
Page
9.5.4 Identifying Exposure Components That Relate to
Vegetation Effects 9-166
9.5.4.1 Role of Concentration 9-166
9.5.4. 2 Role of Duration 9-169
9.5.4.3 Patterns of Exposure 9-169
9.5.4. 4 Frequency of Occurrence of Peak Concentrations 9-172
9.5.4.5 Canopy Structure 9-173
9.5.4. 6 Site and Climate Factors 9-174
9.5.4 . 7 Plant Defense Mechanism - Detoxification 9-174
9.5.5 Ozone Uptake or Effective Dose as an Index 9-174
9.5.5.1 Models of Stomatal Conductance 9-175
9.5.5.2 Nonlinear Response and Developing Flux Indices 9-178
9.5.5.3 Simulation Models 9-179
9.5.6 Summary 9-179
9.6 OZONE EXPOSURE-PLANT RESPONSE RELATIONSHIPS 9-181
9.6.1 Introduction 9-181
9.6.2 Summary of Key Findings/Conclusions from Previous
Criteria Documents 9-182
9.6.3 Ozone Indices and Ambient Exposure 9-193
9.6.4 Effects of Ozone on Annual and Biennial Species 9-212
9.6.4.1 Effects on Growth, Biomass, and Yield of
Individual Species 9-212
9.6.4.2 Effects on Plant Quality 9-217
9.6.4.3 Effects on Foliar Symptoms 9-219
9.6.4.4 Other Effects 9-220
9.6.4.5 Scaling Experimental Data to Field Conditions 9-221
9.6.4.6 European Critical Levels 9-224
9.6.4.7 Summary of Effects on Short-Lived Species 9-227
9.6.5 Effects of Ozone on Long-Lived (Perennial) Species 9-229
9.6.5.1 Herbaceous Perennial Species 9-229
9.6.5.2 Deciduous Woody Species 9-234
9.6.5.3 European Critical Levels 9-240
9.6.5.4 Summary of Effects on Deciduous Woody Species 9-240
9.6.5.5 Evergreen Woody Species 9-241
9.6.5.6 Summary of Effects on Evergreen Woody Species 9-244
9.6.5.7 Scaling Experimental Data to Mature Trees 9-245
9.6.6 Studies with the Chemical EDU 9-249
9.6.7 Summary 9-251
9.7 EFFECTS OF OZONE EXPOSURE ON NATURAL ECOSYSTEMS 9-254
9.7.1 Introduction 9-254
9.7.2 Case Studies 9-256
9.7.2.1 Valley of Mexico 9-256
9.7.2.2 San Bernardino Mountains 9-263
9.7.2.3 Sierra Nevada Mountains 9-268
III-x
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Table of Contents
(cont'd)
Page
9.7.2.4 Appalachian Mountains 9-270
9.7.2.5 Plantago Studies in the United Kingdom 9-271
9.7.2.6 Forest Health in the Carpathian Mountains 9-271
9.7.2.7 Field Exposure System (FACE), Rhinelander,
Wisconsin 9-273
9.7.3 Landscape Condition 9-275
9.7.4 Biotic Condition 9-278
9.7.4.1 Ecosystems and Communities 9-278
9.7.4.2 Species and Populations 9-284
9.7.4.3 Organism Condition 9-287
9.7.5 Ecosystem, Chemical, and Physical Characteristics
(water, soil) 9-295
9.7.5.1 Nutrient Concentrations, Trace Inorganic and
Organic Chemicals 9-295
9.7.6 Ecological Processes 9-296
9.7.6.1 Energy Flow 9-296
9.7.6.2 Material Flow 9-298
9.7.7 Hydrological and Geomorphological 9-301
9.7.8 Natural Disturbance Regimes 9-301
9.7.9 Scaling to Ecosystem Levels 9-302
9.7.9.1 Scaling from Seedlings to Mature Trees 9-303
9.7.9.2 Surveys, Growth Correlations and Stand-Level
Modeling 9-305
9.7.10 Summary of Ecological Effects of Ozone Exposure on
Natural Ecosystems 9-312
9.8 ECONOMIC EVALUATION OF OZONE EFFECTS ON AGRICULTURE,
FORESTRY AND NATURAL ECOSYSTEMS 9-315
9.8.1 Introduction 9-315
9.8.2 The Measurement of Economic Information 9-317
9.8.3 Understanding of Air Pollutants Effects on the Economic
Valuation of Agriculture and Other Vegetation in the
1996 Criteria Document 9-318
9.8.3.1 Agriculture 9-319
9.8.3.2 Forests (Tree Species) and Natural Ecosystems 9-322
9.8.4 Studies Since 1996 of Ozone Exposure Effects on the
Economic Value of Agriculture, Forests, and Ecosystems 9-323
9.8.5 Limitations of Scientific Studies and Economic Information 9-325
9.8.6 Conclusions 9-328
9.9 SUMMARY AND CONCLUSIONS FOR VEGETATION AND
ECOSYSTEM EFFECTS 9-329
9.9.1 Introduction 9-329
9.9.2 Methodology 9-331
9.9.3 Mode-of-Action 9-332
9.9.4 Modification of Growth Response 9-334
Ill-xi
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Table of Contents
(cont'd)
Pas
9.9.5 Exposure Indices 9-337
9.9.6 Ozone Exposure-Plant Response Relationships 9-340
9.9.7 Ecosystem Effects 9-343
9.9.8 Economics 9-345
REFERENCES 9-347
10. TROPOSPHERIC OZONE EFFECTS ON UV-B FLUX AND CLIMATE
CHANGE PROCESSES 10-1
10.1 INTRODUCTION 10-1
10.2 THE ROLE OF TROPOSPHERIC OZONE IN DETERMINING
GROUND-LEVEL UV-B FLUX 10-1
10.2.1 Factors Governing Ultraviolet Radiation Flux at the
Earth's Surface 10-2
10.2.1.1 UV Radiation: Wavelengths, Energies and Depth
of Atmospheric Penetration 10-2
10.2.1.2 Temporal Variations in Solar Flux 10-3
10.2.1.3 Atmospheric Radiative Interactions with Solar
UV Radiation 10-4
10.2.1.4 Modeling Surface UV-B Irradiance 10-10
10.2.2 Factors Governing Human Exposure to Ultraviolet Radiation 10-11
10.2.2.1 Outdoor Activities 10-12
10.2.2.2 Occupation 10-13
10.2.2.3 Age 10-13
10.2.2.4 Gender 10-14
10.2.2.5 Geography 10-15
10.2.2.6 Protective Behavior 10-15
10.2.2.7 Summary of Factors that Affect Human Exposures
to Ultraviolet Radiation 10-16
10.2.3 Factors Governing Human Health Effects due to Ultraviolet
Radiation 10-16
10.2.3.1 Erythema 10-17
10.2.3.2 Skin Cancer 10-19
10.2.3.3 Ultraviolet Radiation Exposure and the Incidence
of Nonmelanoma Skin Cancers 10-20
10.2.3.4 Ocular Effects of Ultraviolet Radiation Exposure 10-26
10.2.3.5 Ultraviolet Radiation and Immune System
Suppression 10-27
10.2.3.6 Protective Effects of Ultraviolet Radiation -
Production of Vitamin D 10-29
10.2.4 Summary and Conclusions for O3 Effects on UV-B Flux 10-30
10.3 TROPOSPHERIC OZONE AND CLIMATE CHANGE 10-31
10.3.1 The Projected Impacts of Global Climate Change 10-32
10.3.2 Solar Energy Transformation and the Components of the
Earth's Climate System 10-36
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Table of Contents
(cont'd)
Page
10.3.3 The Composition of the Atmosphere and the Earth's
Radiative Equilibrium 10-37
10.3.3.1 Forcing of the Earth's Radiative Balance 10-39
10.3.4 Factors Affecting the Magnitude of Climate Forcing by Ozone 10-41
10.3.4.1 Global versus Regional Atmospheric Ozone
Concentrations 10-41
10.3.4.2 Global Versus Regional Atmospheric Ozone Trends .... 10-42
10.3.4.3 The Sensitivity of Ozone-related Forcing Surface
to Albedo 10-45
10.3.5 Estimated Forcing by Tropospheric Ozone 10-45
10.3.5.1 Direct Climate Forcing Due to Ozone 10-45
10.3.5.2 Indirect Forcing Due to Ozone 10-47
10.3.5.3 Predictions for Future Climate Forcing by
Anthropogenic Ozone 10-48
10.3.6 Conclusion 10-48
REFERENCES 10-50
11. EFFECT OF OZONE ON MAN-MADE MATERIALS 11-1
11.11.1 Mechanisms of Ozone Damage and Exposure-Response 11-1
11.11.2 Textiles andFabrics 11-3
11.11.3 Dyes, Pigments, andlnks 11-4
11.11.4 Artists' Pigments 11-5
11.11.5 Surface Coatings 11-8
REFERENCES 11-14
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List of Tables
Number Page
9-1 Advantages and Disadvantages of Protective Chemicals Used in Assessment
of O3 Effects on Plants 9-17
9-2 Advantages and Disadvantages of Bioindicators Used to Study O3 Plant Effects .... 9-18
9-3 Advantages and Disadvantages of Cultivar Comparisons Used in Assessment
of O3 Effects on Plants 9-22
9-4 Advantages and Disadvantages of Various Dendrochronological Techniques
Used in Assessment of O3 Effects on Plants 9-24
9-5 Advantages and Disadvantages of Modeling Techniques Used in Assessment
of O3 Effects on Plants 9-27
9-6 The Flow of Ozone into a Leaf and Possible Reactions 9-39
9-7 Some Rates of Reaction of Ozone With Critical Biochemicals 9-45
9-8 Superoxide Dismutase Isozymes and Isoforms 9-62
9-9 Gene Families and cDNA Clones Used as Probes for SAR (Ward et al, 1991) 9-66
9-10 Proteins Altered by Ozone as Measured by Molecular Biological Techniques
as mRNA Level or Other Gene Activity Rather than Enzyme Activity 9-71
9-11 Interactions Involving O3 and Plant Pathogens 9-97
9-12 Effects of Increased Carbon Dioxide on Ozone-Induced Responses of Plants
at the Metabolic, Physiological, and Whole-Plant Levels 9-130
9-13 Summary of Ozone Exposure Indices Calculated for 3- or 5-Month Growing
Seasons from 1982 to 1991 9-185
9-14 Ozone Exposure Levels (Using Various Indices) Estimated To Cause at Least
10% Crop Loss in 50 and 75% of Experimental Cases 9-187
9-15 SUM06 Levels Associated with 10 and 20% Total Biomass Loss for 50 and
75% of the Seedling Studies 9-190
9-16 Summary of Selected Studies of Ozone Effects on Annual Species 9-195
9-17 Summary of Selected Studies of the Effects of Ozone on Perennial
Herbaceous Plants 9-199
9-18 Summary of Selected Studies of Ozone Effects on Deciduous Trees and Shrubs . . . 9-201
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List of Tables
(cont'd)
Number Page
9-19 Summary of Selected Studies of Ozone Effects on Evergreen Trees and Shrubs . . . 9-206
9-20 Ethylene Diurea Effects on Vegetation Responses to Ozone 9-210
9-21 Ozone Exposures at 35 Rural Sites in the Clean Air Status and Trends Network
in the Central and Eastern United States From 1989 to 1995 9-228
9-22 Essential Ecological Attributes for Natural Ecosystems Affected by O3 9-257
9-23 Case Studies Demonstrating the Ecological Effects of O3 9-264
9-24 The Most Comprehensively Studied Effects of O3 on Natural Ecosystem is
the San Bernardino Mountain Forest Ecosystem 9-266
9-25 Effects of Ozone, Ozone and N Deposition, and Ozone and Drought Stress
on Pinus ponderosa and Pinus jeffreyi in the Sierra Nevada and the San
Bernardino Mountains, California. Citations are Focused on Research
Published Since U.S. EPA (1996) 9-267
9-26 Summary of Responses ofPopulits tremuloides to Elevated CO2 (+200 |imol
mol"1), O3 (1.5 x ambient), or CO2+O3 Compared with Control During Three
Years of Treatments at the Aspen FACE Project (Modified from Karnosky
et al. 2003a) 9-274
10-1 Examples of Impacts Resulting From Proj ected Changes in Extreme
Climate Events 10-34
10-2 CTM Studies Assessed by the IPCC for its Estimate of the Change in Global,
Total Column O3 Since the Pre-Industrial Era 10-44
10-3 Tropospheric O3 Change (HO3) in Dobson Units (DU) Since Pre-industrial
Times, and the Accompanying net (SW plus LW) Radiative Forcings (Wirf2),
After Accounting for Stratospheric Temperature Adjustment (using the Fixed
Dynamical Heating method) 10-46
11-1 Average 24-h Ozone Concentrations Producing the Highest Frequency of
Cracks of a Certain Length in the Middle and Central Zones of the Rubber
Test Strips 11-3
11-2 Cuprammonium Fluidity of Moist Cotton Cloth Exposed to 20 to 60 ppb Ozone .... 11-4
11-3 Color Change After 12 Weeks of Exposure to a Mixture of Photochemical
Oxidants 11-11
III-xv
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List of Figures
Number Page
9-1 Ozone uptake from the atmosphere 9-30
9-2 Absorption and transformation of O3 within the leaf 9-32
9-3 The uptake of O3 into the leaf. Each of the individual concentration layers of
O3 represents a different process of movement and of plant/ microenvironmental
interaction 9-36
9-4 The microarchitecture of a dicot leaf 9-37
9-5 The change in the O3 concentration inside a leaf with time 9-38
9-6 Possible transformations of O3 within a leaf 9-41
9-7 Possible reactions of O3 within water 9-41
9-8a,b The Crigee mechanism of O3 attack of a double bond 9-44
9-9 Varied ESR radicals, trapped and not, generated by ozone under somewhat
physiological conditions 9-47
9-10 Pathogen-Induced Hypersensivity 9-50
9-11 The interaction of H2O2 and Ca2+ movements with ABA-induced
stomatal closure 9-52
9-12 The reaction of ascorbate within the apoplasm of the cell wall and its ultimate
reduction/oxidations 9-56
9-13 The pathway leading from phospholipids to jasmonic and traumatic acid 9-70
9-14 The production of Rubisco and its Calvin Cycle pathway reactions 9-75
9-15 Linkage of senescence with hypersensitivity reactions and first event of O3
attack of plants 9-82
9-16 Diagrammatic representation of several exposure indices, illustrating how they
weight concentration and accumulate exposure 9-162
9-17 Trends in May-September 12-h SUM06, peak 1-h ozone concentration and
number of daily exceedances of 95 ppb for Crestline in 1963-1999 in relation
to trends in mean daily maximum temperature for Crestline and daily reactive
organic gases (ROG) and oxides of nitrogen (NOX) for San Bernardino county .... 9-168
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List of Figures
(cont'd)
Number Page
9-18 Distribution of biomass loss predictions from Weibull and linear
exposure-response models that relate biomass to O3 exposure 9-188
9-19 A conceptual diagram of processes and storage pools in sources and sinks that
are affected by ozone exposure 9-276
9-20 Common anthropogenic stressors and the essential ecological attributes
they affect 9-313
10-1 Extraterrestrial solar flux measured by the satellite UARS SOLSTICE
instrument (dotted line) 10-3
10-2 Ozone column abundances from the years 1990 to 1992 for 0, 40, and 80° N
as well as 80° S 10-6
10-3 The sensitivity of ground-level UV flux to a 1 DU change in total column
O3, under clear sky conditions, as a function of solar zenith angle (SZA) 10-10
10-4 Complexity of factors that determine human exposure to UV radiation 10-11
10-5 Estimated global mean radiative forcing exerted by gas and various particle
phase species for the year 2000, relative to 1750 10-40
10-6 Mid-tropospheric O3 abundance (ppb) in northern midlatitudes (36 °N-59 °N)
for the years 1970 to 1996 10-43
11-1 In-service fading of nylon 6 yarn inside house 11-6
11-2 In-service fading of nylon 6 yarn outside house 11-7
11-3 Observed color changes for natural colorant-on-paper systems during exposure
to 0.40 ppm ozone at 25 °C ± 1 °C, 50% RH, in the absence of light 11-9
11-4 Observed color changes for natural colorant-on-site during exposure to
0.40 ppm ozone at 25 °C ± 1 °C, 50% RH, in the absence of light 11-10
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Authors, Contributors, and Reviewers
To be inserted in Second External Review draft.
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U.S. ENVIRONMENTAL PROTECTION AGENCY
PROJECT TEAM FOR DEVELOPMENT OF AIR QUALITY CRITERIA
FOR OZONE
To be inserted in Second External Review draft.
-------
U.S. ENVIRONMENTAL PROTECTION AGENCY
SCIENCE ADVISORY BOARD (SAB) STAFF OFFICE
CLEAN AIR SCIENTIFIC ADVISORY COMMITTEE (CASAC) OZONE
REVIEW PANEL*
To be inserted in Second External Review draft.
III-xx
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Abbreviations and Acronyms
AQCD
CxT
CI
CO2
COP
CTM
DNA
DNA
EPA
GSH
H+
H202
HFC
HO
HO2
IPCC
IR
mRNA
NHAPS
NMHC
NO2
NOX
NRC
0('D)
03
OH
PAN
Air Quality Criteria Document
Concentration times duration of exposure
color index
carbon dioxide
Conference of Parties
Chemistry Transport Model
Deoxyribonucleic acid
Deoxyribonucleic acid
U.S. Environmental Protection Agency
Glutathione
hydrogen ion
hydrogen peroxide
hydrofluorocarbon
hydroxyl
hydroperoxyl; hydroperoxy
Intergovernmental Panel on Climate
infrared radiation
Messenger ribonucleic acid
Change
National Human Activity Pattern Survey
nonmethane hydrocarbon
nitrogen dioxide
nitrogen oxides
National Research Council
electronically excited oxygen atom
ozone
hydroxyl; hydroxy
Peroxyacetyl nitrate
-------
RNA Ribonucleic acid
RO2 organic peroxyl; organic peroxy
SO2 sulfur dioxide
SOD Superoxide dismutase
SUMOO Sum of all hourly average concentrations
SUM06 Seasonal sum of all hourly average concentrations > 0.06 ppm
TAR Third Assessment Report
UNEP United Nations Environment Program
UV Ultraviolet
UV-A
UV-B Ultraviolet radiation of wavelengths from 280 to 320 nm
uv-c
W126 cumulative integrated exposure index with a sigmoidal weighting function
WMO World Meteorological Organization
ZAPS Zonal Air Pollution System
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i 9. ENVIRONMENTAL EFFECTS: OZONE EFFECTS
2 ON VEGETATION AND ECOSYSTEMS
3
4
5 9.1 INTRODUCTION
6 The preceding chapters of this document focused on discussion of: (a) background
7 information regarding ozone (O3)-related atmospheric chemistry, air quality and exposure
8 aspects; and (b) dosimetric/health effects aspects, as well as the integrative synthesis of key
9 information drawn from such chapters of most importance for EPA review of the primary O3
10 NAAQS. On the other hand, this and the next two chapters assess available scientific
11 information on O3-related welfare effects of most pertinence for the review of secondary O3
12 NAAQS. This includes discussion of three classes of environmental effects:
13 • O3 effects on vegetation and ecosystem (Chapter 9);
14 • the role of tropospheric O3 in climate change, including determining of ground-level
UV-B flux (Chapter 10); and
15 • O3 effects on manmade materials (Chapter 11).
16 As for the organizational structure of this chapter, after this opening Introduction,
17 Section 9.2 highlights key features of methodologies used in vegetation research, followed by
18 discussion of vegetation species response/mode-of-action aspects (Section 9.3) and factors that
19 modify functional and growth responses (Section 9.4). Effects-based air quality exposure and
20 dose-response indices are next discussed in Section 9.5, followed by assessment of O3 exposure-
21 plant response relationships (Section 9.6) and effects on natural ecosystems (Section 9.7). The
22 economic evaluation of O3 effects on agriculture, forestry, and natural ecosystems is then
23 discussed in Section 9.8. Lastly, the chapter closes with Section 9.9, an overall summary of key
24 findings and conclusions.
25
26
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1 9.2 METHODOLOGIES USED IN VEGETATION RESEARCH
2 9.2.1 Introduction
3 The scale of investigations into the direct effects of O3 on plants range from evaluation of
4 sub-cellular to cellular, organismal, population, community, and ecosystem responses, each with
5 its particular experimental methodologies and suite of specialized instrumentation, equipment,
6 facilities, and experimental protocols. These investigations generate data; whereas other types of
7 methodologies exist for the handling of data and statistical analysis as well as the utilization of
8 data in developing the different exposure metrics or indices used to define exposure, quantitative
9 exposure-response relationships, and computer simulation models of these. The objective of this
10 section is not to provide an updated encyclopedia of all the methods that have been used, but
11 rather to focus on approaches that have:
12 (1) led to improved understanding of the quantitatively measurable growth and
development responses of plants and plant communities to O3, or
13 (2) provided information about the extent and geographic distribution of the
responses of herbaceous and woody plants, both cultivated and native, to
ambient O3 exposures.
14 The first objective is essential for determining dose-response functions used in the development
15 of impact and risk assessments of the effects of O3 and usually involves treating plants to a range
16 of artificial O3 exposures. The second objective is essential for determining the geographic
17 distribution of the risk and usually involves plants subjected to ambient air O3 exposures.
18 The types of methodologies used by biochemists, molecular biologists, or plant
19 physiologists, whose interests lie in determining effects on specific constituents or in
20 understanding the mode of action of O3, are not discussed here. Methods used to characterize
21 the O3 content of ambient air and to define exposure and exposure-response relations are
22 discussed in Sections 9.5 and 9.6, respectively.
23 The methodologies for exposure-response studies have used many different types of
24 exposure facilities and protocols and have employed a range of statistical approaches to the
25 analysis and interpretation of data. Most of the studies have been conducted using major
26 agricultural crop species. The methodologies have improved over the years as a result of the
27 development, availability, or application of new or improved instrumentation, physical systems,
28 and numerical approaches to data analysis. Yet equally important to the roles played by these
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1 advances has been the clearer understanding that has emerged from earlier work identifying the
2 type of experimentation needed to achieve realistic assessments of the magnitude and extent of
3 the impact of O3 on vegetation of all types. As a result, significantly increased attention is now
4 being paid to field observations of responses to ambient O3 pollution, particularly to the
5 responses of forest trees and native vegetation.
6 Other than in various exploratory studies that have used chamber-based steady-state
7 exposure concentrations (so-called "square-wave" exposures), the trend in experimental
8 exposure protocols has been to attempt to expose plants under conditions as natural as possible
9 to temporal profiles that simulate the real-world, either by conducting experiments in the field or
10 in elaborately controlled environment facilities that provide simulated field conditions.
11 Previous Air Quality Criteria Documents for Ozone and Other Photochemical Oxidants
12 (U.S. Environmental Protection Agency, 1986; 1996) described the time-course for these
13 methodological developments. Although this section provides a brief overview of the
14 methodologies used in the past, and their limitations, it focuses mainly on those techniques that
15 have come into prominence over the last decade. This has been aided considerably by several
16 compilations of experimental methodologies and facilities, such as the earlier comprehensive
17 review for the U.S. Environmental Protection Agency/National Acid Precipitation Assessment
18 Program by Hogsett et al. (1987a; 1987b) and the more recent reviews by Manning and Krupa
19 (1992), Musselman and Hale (1997) and Karnosky et al. (2001).
20
21 9.2.2 Methods Involving Experimental Exposures to O3
22 9.2.2A "Indoor", Controlled Environment, and Greenhouse Chambers
23 The earliest experimental investigations of the effects of O3 on plants utilized simple glass
24 or plastic-covered chambers, often located within greenhouses, into which a flow of O3-enriched
25 air or oxygen could be passed to provide the exposure. The types, shapes, styles, materials of
26 construction, and locations of these chambers were as numerous as the different investigators
27 and, in spite of providing little resemblance to real-world conditions, they yielded much of the
28 basic information on the visible and physiological effects on plants. The construction and
29 performance of more elaborate and better instrumented chambers dating back to the 1960s has
30 been well-summarized in Hogsett et al. (1987a), including those installed in greenhouses (with
31 or without some control of temperature and light intensity).
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1 One greenhouse chamber approach that continues to yield useful information on the
2 relationships of O3 uptake to both physiological and growth effects employs continuous stirred
3 tank reactors (CSTRs) first described by Heck et al. (1978). Although originally developed to
4 permit mass-balance studies of O3 flux to plants, their use has more recently widened to include
5 short-term physiological and growth studies of O3 x CO2 interactions (e.g., Costa et al., 2001;
6 Heagle et al., 1994b; Loats and Rebbeck, 1999; Rao et al.. 1995; Reinert and Ho. 1995; and
7 Reinert et al.. 1997), and of surveys of native plant responses to O3 (Orendovici et al., 2003).
8 In many cases, supplementary lighting and temperature control of the surrounding structure have
9 been used to control or modify the environmental conditions, e.g., Heagle et al. (1994a).
10 Many investigations have utilized commercially available controlled environment
11 chambers and walk-in rooms adapted to permit the introduction of a flow of O3 into the
12 controlled air-volume. Such chambers continue to find use in genetic screening and in
13 physiological and biochemical studies aimed primarily at improving our understanding of mode
14 of action. For example, some of the ongoing studies of the O3 responses ofPlantago major
15 populations have been conducted in controlled environment chambers (Reiling and Davison,
16 1994; Whitfield et al., 1996b).
17 The environmental conditions provided by indoor chambers of any type will always
18 preclude the use of the information obtained with such chambers in predicting O3 effects in the
19 natural environment, because the environmental conditions will always be measurably different
20 from field conditions. However, highly sophisticated controlled environment chambers such as
21 those described by Langebartels et al. (1997), which are subdivided into aerial and root
22 compartments, with dynamic control of light intensity and photoperiod, air and soil temperature,
23 humidity, soil moisture, wind speed, and exposure to O3, may come close to simulating specific
24 natural conditions. Such chambers have provided meaningful insights into a wide array of early
25 biochemical responses of plants to O3. They can minimize confounding factors that make indoor
26 chamber studies only rarely able to be extrapolated to field conditions, i.e., that shoots and roots
27 develop under different temperature regimes.
28 The applicability of the results of many chamber studies is also limited by their use of
29 container-grown plants. Several recent studies have raised serious questions as to the relevance
30 of pot-based studies to the true field situation (even when the studies are conducted in the field.)
31 Most of the questions have concerned studies of the effects of CO2 enrichment, as discussed in
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1 Section 9.4.7.1, but are also relevant to O3 enrichment studies, as shown by Whitfield et al.
2 (1996a). Whitfield et al., reported significant interactive effects between O3 and soil volume on
3 the growth ofPlantago major. They noted that although container size may limit root and,
4 hence, plant growth, the reverse may also be true for single plants in large containers which do
5 not experience typical field competition for resources.
6
7 9.2.2.2 Field Chambers
8 Although closed field chambers have largely fallen out of favor in recent years, closed
9 "Solardome" field chambers (Lucas et al., 1987; Rafarel and Ashenden, 1991) have been used
10 recently in studies of O3 x acid mist interactions (Ashenden et al., 1995; 1996).
11 Concern over the need to establish realistic plant-litter-soil relationships as a prerequisite to
12 studies of the effects of O3 and CO2 enrichment on ponderosa pine (Pinus ponder osa) seedlings
13 led Tingey et al. (1996) to develop closed, partially environmentally controlled, sun-lit chambers
14 ("terracosms") incorporating 1 m-deep lysimeters containing forest soil that retained the
15 appropriate horizon structure.
16 In general, field chamber studies are dominated by the use of various versions of the open-
17 top chamber (OTC) design, first described by Heagle et al. (1973) and Mandl et al. in 1973.
18 Most chambers are ~3 m in diameter with 2.5 m high walls. Hogsett et al. (1987a) described in
19 detail many of the various modifications to the original OTC designs that appeared subsequently,
20 e.g., the use of larger chambers to permit exposing small trees (Kats et al., 1985); and grapevines
21 (Mandl et al., 1989), the addition of a conical baffle at the top to improve ventilation (Kats et al.,
22 1976), a frustrum at the top to reduce ambient air incursions, and a plastic rain-cap to exclude
23 precipitation (Hogsett et al., 1985). All of these modifications included the discharge of air via
24 ports in annular ducting or interiorly perforated double-layered walls at the base of the chambers
25 to provide turbulent mixing and upward mass flow of air.
26 Wiltshire et al. (1992) described a large open-top chamber suitable for small trees with
27 roll-up sides that permitted the trees to be readily subjected from time to time to episodic,
28 normal, "chamberless" environmental conditions. In the 6 m-high OTCs described by Seufert
29 and Arndt (1985) used with Norway spruce (Picea abies) trees, a second zone of annular
30 enrichment was also provided between 4 and 5 m. The use of OTCs was adopted for the large
31 European Stress Physiology and Climate Experiment on the effects of CO2 and O3 on spring
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1 wheat (ESPACE-wheat), conducted in 1994-1996 at field sites in eight countries (Jager et al.,
2 1999). However, typical European chambers have the introduction of O3-enriched air at or
3 above canopy height. The relatively low costs of fabrication, operation and maintenance has
4 favored OTC use in field studies (Fangmeier et al., 1992; Musselman and Hale, 1997). The air
5 supplied to the chambers can be readily filtered through activated charcoal to reduce the O3
6 concentration or it can be enriched with O3 to provide a range of exposures.
7 All field chambers create internal environments that differ from ambient air, giving rise to
8 so-called "chamber effects" with the modification of microclimatic variables, including reduced
9 and uneven light intensity, uneven rainfall, constant wind speed, reduced dew formation, and
10 increased air temperatures (Fuhrer, 1994; Manning and Krupa, 1992). Several shortcomings of
11 the OTC design and operation relate to the means of introduction and mixing of enriched air to
12 produce a definable exposure. First, the plants are subjected to constant turbulence, which,
13 through increased uptake resulting from the consequently low boundary layer resistance to
14 diffusion, may lead to overestimation of the magnitude of real-world cause-effect relationships
15 (Krupa et al., 1995; Legge et al., 1995). Second, the introduction of the O3-enriched air in the
16 lower part of chambers described by Heagle et al. (1973) and Mandl et al. (1973) results in a O3
17 concentration gradient that decreases with increasing height, the converse of the situation
18 observed in ambient air, in which the O3 concentration decreases markedly from above a plant
19 canopy to ground level (Griinhage and Jager, 1994; Pleijel et al., 1995; Pleijel et al., 1996).
20 Concern that studies conducted in such OTCs may somewhat overestimate the effects of O3 led
21 to the European design which provides a decreasing downward gradient. It seems unlikely that
22 the "chamber effects" produced by the two designs will be the same. These issues are discussed
23 more fully in Section 9.2.2.4.
24 It should also be noted that, although originally developed for exposing row drops in the
25 field, many recent studies employing OTCs have used potted plants in order to include or control
26 edaphic or nutritional factors or water relations within the experimental design. Therefore, the
27 same caveats as those discussed above (Section 9.2.2.1) with regard to extrapolating results of
28 pot studies to true field conditions apply to OTC studies.
29 The difficulties faced in the experimental exposure of forest trees to air pollutants in
30 chambers (e.g., Seufert and Arndt [1985]) led to the development of branch chambers such as
31 those described by Ennis et al. (1990), Houpis et al. (1991) and Teskey et al. (1991). These
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1 chambers are essentially large cuvettes and, as noted by Musselman and Hale (1997), share
2 many of the characteristics of CSTRs, i.e., transparent walls, internal fans, and inlet and outlet
3 monitors to permit the determination of O3 uptake, CO2 exchange, and transpiration. Although
4 they make it possible to expose whole branches to different O3 regimes, the relevance of the data
5 they yield in regard to the whole tree may be questioned. As noted by Saxe et al. (1998), the
6 inevitable change in environmental conditions resulting from the isolation of the branch may
7 cause different responses from those that would be obtained if the whole tree was subjected to
8 the same environmental conditions.
9
10 9.2.2.3 Plume Systems
11 Plume systems are chamberless exposure facilities in which the atmosphere surrounding
12 plants in the field is modified by the injection of pollutant gas into the air above or around them
13 from multiple orifices that are spaced to permit diffusion and turbulence so as to establish
14 relatively homogeneous conditions as the individual plumes disperse and mix with the ambient
15 air. As pointed out by Manning and Krupa (1992), they can only be used to increase the O3
16 levels in the ambient air. The volume of air to be modified is unconfined, and three approaches
17 have been used to achieve desired pollutant concentrations in the air passing over the plants
18 producing various systems that:
19 (1) achieve a concentration gradient, in most instances dependent upon the direction of
the prevailing wind;
20 (2) achieve spatially uniform concentrations over a plot, dependent upon wind
direction; and
21 (3) seek to achieve spatially uniform concentrations over a plot, independent of wind
speed and direction.
22 Gradient systems created by dispensing a pollutant gas into the air at canopy level from
23 perforated horizontal pipes arranged at right angles to the prevailing wind were described for
24 SO2 studies in the early 1980s. A modified gradient system for O3 was used by Bytnerowicz
25 et al. (1988) to study effects on desert species, but there appear to have been no recent
26 applications of the method. A gradient O3-exclusion system is discussed in Section 9.2.3.1.
27 Systems designed to achieve spatially uniform pollutant levels by ensuring that the release
28 of a pollutant is always on the upwind side of the study site were also originally described for
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1 SO2 studies, e.g. Greenwood et al. (1982). However, the adaptation of these concepts as
2 introduced by McLeod et al. (1985) in constructing a large circular field site for exposing crops
3 to SO2 led to the subsequent development of both the large-scale O3 and SO2 fumigation system
4 for forest trees in the United Kingdom in 1985 (the Liphook Forest Fumigation Project; McLeod
5 et al., 1992), the smaller system for O3 fumigation constructed at Kuopio, Finland in 1990
6 (Wulff et al., 1992), and the free-air carbon-dioxide enrichment (FACE) systems of gas dispersal
7 over crops (Hendrey and Kimball, 1994) and forest trees (Hendrey et al., 1999). Although
8 originally designed to provide chamberless field facilities for studying the CO2 effects of climate
9 change, large forest tree FACE systems have recently been adapted to include the dispensing of
10 O3 (Karnosky et al., 1999). Volk et al. (2003) have recently described a system for exposing
11 grasslands that uses 7-m diameter plots. FACE systems discharge the pollutant gas (and/or CO2)
12 through spaced orifices along an annular ring (or torus) or at different heights on a ring of
13 vertical pipes. Computer-controlled feedback from the monitoring of gas concentration
14 regulates the feed rate of enriched air to the dispersion pipes. Feedback of wind speed and
15 direction information ensures that the discharges only occur upwind of the treatment plots, and
16 that discharge is restricted or closed down during periods of low wind speed or calm conditions.
17 The diameter of the arrays and their heights in some FACE systems (25-30 m) requires large
18 throughputs of enriched air per plot, particularly in forest tree systems. The cost of the
19 throughputs tends to limit the number of enrichment treatments, although Hendrey et al. (1999)
20 have argued that the cost on an enriched volume basis is comparable to that of chamber systems.
21 An alternative to the FACE system to free-air fumigation uses a horizontal grid system
22 through which pollutant-air enriched is discharged over the canopies of plants in field plots. The
23 original design, termed the Zonal Air Pollution System (ZAPS), was developed for studying the
24 effects of SO2 on native grasslands (Lee et al., 1975), and it was later modified by Runeckles
25 et al. (1990) by randomly dividing each of three treatment plots into four sub-plots, each with
26 different numbers of discharge orifices to provide various levels of O3 enrichment. With the
27 ZAPS system, changes in wind direction and speed result in varying degrees of carry-over from
28 sub-plot to sub-plot, effectively resulting in twelve stochastically different seasonal exposures.
29 The system was used for studies of growth effects on field crops and 2- to 4-year old Douglas fir
30 (Pseudotsuga menziesii) saplings (Runeckles and Wright, 1996).
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1 A larger ZAPS design was used by Wilbourn et al. (1995) on a grass (Loliumperenne)-
2 clover (Trifolium repens) mixture and by Ollerenshaw et al. (1999) on oilseed rape (Brassica
3 napus), whereby four replicate field plots were exposed to intermittent constant additions of O3
4 to ambient air. A ZAPS design with eight spatially separated treatment plots was also developed
5 to obtain crop response data used in assessing crop losses in the Fraser Valley, British Columbia,
6 Canada (Runeckles and Bowen, 2000).
7 The FACE-type facility developed for the Kranzberg Ozone Fumigation Experiment in
8 Germany that began in 2000 (KROFEX; Werner and Fabian, 2002; Nunn et al., 2002) to study
9 the effects of O3 on mature stands of beech (Fagus sylvaticd) and spruce (Picea abies) trees is
10 more truly a zonal system that functions independently of wind direction. The enrichment of a
11 large volume of the ambient air immediately above the canopy takes place via orifices in vertical
12 tubes suspended from a horizontal grid supported above the canopy.
13 Recognizing the difficulties of modifying the aerial environments of large trees, Tjoelker
14 et al. (1994) devised a free-air system for exposing branches of sugar maple (Acer saccharum)
15 trees to O3. Near the ends of up to 10 branches, enriched air was discharged through small holes
16 in 38-cm-diameter loops of 0.635-cm-OD teflon tubes positioned 20-30 cm below the terminal
17 foliage cluster.
18 Although plume systems make virtually none of the modifications to the physical
19 environment that are inevitable with chambers, their successful use depends on selecting the
20 appropriate numbers, sizes, and orientations of the discharge orifices to avoid "hot-spots"
21 resulting from the direct impingement of jets of pollutant-enriched air on plant foliage (Werner
22 and Fabian, 2002). However, because mixing is unassisted and completely dependent on wind
23 turbulence and diffusion, local gradients are inevitable even in large-scale FACE systems. Both
24 FACE and ZAPS systems have provisions for shutting down under low wind speed or calm
25 conditions and for an experimental area that is usually defined within a generous border, in order
26 to strive for homogeneity of the exposure concentrations within the treatment area. They are
27 also both dependent upon continuous computer-controlled feedback of the O3 concentrations in
28 the mixed treated air and the meteorological conditions.
29
30
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1 9.2.2.4 Comparative Studies
2 All experimental approaches to the exposure of plants to O3 have shortcomings. The use of
3 laboratory, greenhouse, or field chambers raises concerns for the roles of "chamber effects" on
4 micrometeorology as well as the consistent turbulence over and within the plant canopy during
5 chamber operation, in modifying plant responses.
6 With the widespread use of the OTC, numerous studies have demonstrated increased air
7 temperatures and decreased light intensities as regularly noted features of their use. However,
8 the question to be answered is whether or not these differences affect plant response to O3.
9 As noted in the 1996 criteria document (U.S. Environmental Protection Agency, 1996), evidence
10 from the comparative studies of OTCs and from closed chamber and O3-exclusion exposure
11 systems on the growth of alfalfa (Medicago saliva) by Olszyk et al. (1986a) suggested that, since
12 significant differences were found for fewer than 10% of the growth parameters measured, the
13 responses were, in general, essentially the same regardless of exposure system used and
14 "chamber effects" did not significantly affect response. In 1988, Heagle et al. (1988) concluded:
15 "Although chamber effects on yield are common, there are no results showing that this will
16 result in a changed yield response to O3." Several more recent studies have, however, indicated
17 that the temperature effect alone may be sufficient to cause significant (though modest) shifts in
18 quantitative responses to O3, both positive and negative, as discussed below in Section 9.4.3.2.
19 Still, it is not clear whether these effects are directly related to temperature or are the result of
20 temperature interactions with other environmental variables. For example, Olszyk et al. (1992)
21 undertook a 3-year study of the impact of O3 on Valencia orange trees in large OTCs to
22 determine if "insidious differences in microclimatic conditions could alter plant growth
23 responses and susceptibility to pollutant stress." Non-filtered chambers were found to have
24 somewhat lower average O3 concentrations than the ambient air, with fewer hourly exceedances
25 of 100 ppb. In cool seasons, stomatal conductance was also lower, implying lower O3 uptake.
26 However, the cumulative fruit yields were doubled in the chamber trees even though
27 photosynthetically active radiation was consistently reduced by about 19%, while leaf
28 temperatures averaged more than 2°C higher. These data may be somewhat extreme, but they
29 emphasize the need to avoid automatically assuming that OTCs yield response data that are all
30 immediately relevant to the real world, particularly since, as in this study, no O3 enrichment was
31 involved as a complicating factor.
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1 Plume systems avoid "chamber effects," but because they rely solely upon diffusion and
2 natural turbulence to modify the ambient O3 concentration, they may fail to achieve homogeneity
3 of the air to which the plants are exposed and may give rise to "hot spots" in which the enriched
4 air jets are inadequately diluted and impinge directly on foliage. A further deterrent to their
5 widespread use is the large-scale generation of O3 needed, which has in most cases limited the
6 numbers of treatments that can be included in an experimental design. In spite of the various
7 advantages and disadvantages of the two field systems, it appears that no direct comparative
8 studies have been conducted.
9
10 9.2.2.5 Ozone Generation Systems
11 Two approaches have been used to generate the O3 needed for enrichment purposes from
12 air or oxygen: (1) high voltage static discharge and (2) high intensity UV-irradiation. Using
13 gaseous oxygen as feedstock, both generate O3-enriched oxygen, free from other impurities.
14 However, the use of high voltage discharge equipment with air as feedstock requires that the
15 output be scrubbed with water to remove appreciable amounts of the higher oxides of nitrogen
16 (especially N2O5, nitric acid vapor) that form concurrently with O3 (Brown and Roberts, 1988;
17 Taylor et al., 1993).
18
19 9.2.2.6 Experimental Exposure Protocols
20 A few recent chamber studies of physiological or biochemical effects have continued to
21 use "square-wave" exposure profiles typified by a rapid rise to and falling off from a steady
22 target concentration. However, most recent experimentations into O3 effects on plant growth and
23 development have employed either simulations of the diurnal ambient O3 profile or
24 enhancement/reduction of the ambient O3 concentrations.
25 Although Hogsett et al. (1985), Lefohn et al. (1986), and others have described the use in
26 controlled chambers of daily exposure profiles based on observed ambient O3 profiles and such
27 profiles were used in the elaborately controlled chamber studies of Langbartels et al. (1997),
28 several recent chamber studies have used simpler computer-controlled half- or full-cosine wave
29 profiles to simulate the typical daily rise and fall in ambient O3 levels (McKee et al., 1997a;
30 1997b; Mazarura, 1997).
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1 The early studies with OTCs involved adding constant levels of O3 to ambient air O3
2 concentrations, but all recent studies have used enrichment delivery systems that maintain
3 proportionality to and track ambient O3 concentrations to produce levels that more closely
4 resemble field observations. Both FACE and ZAPS studies have used proportional enrichment
5 to provide a range of treatments, although Wilbourn et al. (1995) and Ollerenshaw et al. (1999)
6 adjusted their systems manually to obtain a relatively constant target concentration during
7 exposure episodes.
8
9 9.2.3 Methods Involving Exposures to O3 in Ambient Air
10 The experimental methods discussed above are largely aimed at developing quantitative
11 growth-response functions to permit the estimation of the effects of different ambient O3
12 scenarios. Because such methodologies usually involve exposures to higher than ambient O3
13 levels, the applicability of the functions obtained may, to some extent, be relevant only to
14 locations that are naturally subjected to high ambient O3 levels. Furthermore, as pointed out in
15 Section 9.4, the response functions that they generate rarely incorporate other environmental,
16 genetic, and physiological factors, many of which can severely modify the magnitude of the
17 response to O3. The consequences of ignoring such modifications have been well stated by De
18 Santis (1999). The European standard for protecting crops (based on the AOT40 index) was
19 derived from OTC studies of O3-induced grain loss of wheat observed in experiments conducted
20 mostly in non-Mediterranean locations. However, the impact of ambient O3 on wheat yields in
21 the Po Valley of northern Italy is much less than the devastatingly high loss (> 60%) suggested
22 by the seasonal exceedances of the standard. On a similar note, Manning (2003) has recently
23 urged the absolute necessity of seeking "ground truth" as verification of the nature and
24 magnitude of impacts on vegetation suggested by response functions using ambient O3
25 monitoring data.
26 Such concerns clearly show that attention needs to be focused on incorporating
27 consideration of environmental and other factors into the response functions upon which
28 standards are based. This will require the development of improved simulation response models.
29 These concerns have also led to increasing attention being paid to seeking and developing
30 alternative approaches to the assessment of impact and the geographic extent of such impact,
31 approaches that are based on in situ exposures to ambient or sub-ambient O3 levels. Although
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1 one approach, the use of air exclusion systems, requires experimental facilities, the other
2 approaches are generally based on simple field observations or measurements and, hence, can be
3 undertaken on a wide geographic scale.
4
5 9.2.3.1 Air-exclusion Systems
6 The term, air-exclusion system, usually refers to chamberless field systems specifically
7 designed to protect plants from exposure to polluted air by blowing filtered air through their
8 canopies. Hogsett et al. (1987a; 1987b) describe several dedicated systems developed in the
9 1960s and 1970s, but there appear to have been no recent O3-exclusion studies using systems
10 specifically designed for the purpose since those described by Olszyk et al. (1986a; 1986b).
11 Their system, a modification of the earlier system of Jones et al. (1977), consisted of perforated
12 31.8-cm OD inflatable polyethylene tubes laid between crop rows and supplied with charcoal -
13 filtered air. By increasing the size of the orifices progressively in sections along the 9-m length
14 of the tubes, an exclusion gradient was created with a progressive decrease in O3 levels in the air
15 surrounding the crop from one end of the system to the other. The system was used for studies
16 on alfalfa (Medicago sativaL.) comparing plant response in OTCs, closed field chambers, the
17 air-exclusion system, and ambient air plots (as discussed above in Section 9.2.2.4).
18 An air-exclusion component has also been part of the overall design of the many OTC
19 experiments. Charcoal-filtered air or mixtures of charcoal-filtered and ambient air to chambers
20 as part of the overall design.
21
22 9.2.3.2 Natural Gradients
23 Naturally occurring locational differences in ambient O3 concentrations hold potential for
24 the examination of plant response along a gradient of such concentrations. However, few such
25 gradients can be found which meet the rigorous statistical requirements for comparable site
26 characteristics such as soil type, temperature, rainfall, radiation, and aspect (Manning and Krupa,
27 1992), although with small plants, soil variability can be avoided by the use of potted plants.
28 Studies in the 1970s used the natural gradients occurring in Southern California to assess
29 yield losses of alfalfa (saliva) and tomato (Lycopersicon esculentum L.} (Oshima et al., 1976;
30 1977). A transect study of the impact of O3 on the growth of clover (Trifolium repens L.} and
31 barley (Hordeum vulgar e L.) in the United Kingdom was confounded by differences in the
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1 concurrent gradients of SO2 and NO2 pollution (Ashmore et al., 1988). Studies of forest tree
2 species in national parks in the eastern U.S. (Winner et al., 1989) revealed increasing gradients
3 of O3 and visible foliar injury with increased elevation.
4
5 9.2.3.3 Use of Chemical Protectants
6 The use of protective chemicals is a relatively inexpensive, promising alternative to
7 experimental field exposures in chambers or free-air systems for determining plant response to
8 O3. Several chemical compounds (antioxidants, antisenescence agents, fungicides, pesticides,
9 etc.) have been known for many years to provide plants some protection from photochemical
10 oxidants, such as O3 (Manning and Krupa, 1992). Most of these chemicals were originally used
11 as a one-time application to reduce visible injury caused by acute O3 exposures. The most
12 widely used and popular of these has been ethylenediurea (EDU). Carnahan et al. (1978)
13 reported that EDU protected pinto bean (Phaseolus vulgaris) from acute O3 injury. After this
14 initial investigation, EDU was shown to suppress visible O3 injury on several species of plants
15 under both controlled and field conditions (Brennan et al., 1987; Clarke et al., 1983). However,
16 due to lack of a commercial market for this product, its commercial manufacture was largely
17 discontinued. Other chemicals, including benomyl (Manning et al., 1974), carboxin (Rich et al.,
18 1974), ascorbic acid (Dass and Weaver, 1968) and others (Manning and Krupa, 1992), exhibited
19 some beneficial effects in reducing visible O3 injury.
20 Despite their widespread use for O3-effects studies, OTCs are not effective in determining
21 plant responses in the field under truly natural conditions, as discussed above, and are not
22 efficient or easily used in unmanaged ecosystems such as wilderness areas (Manning and Krupa,
23 1992). For these reasons, a renewed interest in using protective chemicals as research tools
24 occurred in the early 1990s. These chemicals (mostly EDU) have recently been used in studies
25 of different plant species, both in the United States (Bergweiler and Manning, 1999; Kuehler and
26 Flagler, 1999) and in Europe (Bortier et al., 2001b; Pleijel et al., 1999; Wu and Tiedemann,
27 2002), to determine if ambient O3 concentrations affect plant growth and productivity or are just
28 exacerbating foliar injury. For example, Bortier et al. (2001a) injected seedlings of an O3-
29 sensitive poplar (Populus nigra) clone with EDU and measured growth over a 1-year period at a
30 field site near Brussels, Belgium. Over the growing season, stem diameter increment was
31 significantly higher (16%), biomass was increased (9%), and foliar O3 symptoms were slightly
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1 less for the EDU-treated seedlings. Ozone levels were reported to be low (AOT40 = 6170 ppb.h,
2 May-September) during the exposure period. In another study, Manning et al. (2003) applied
3 EDU (foliar spray) and sodium erythorbate (NaE) at various concentrations, biweekly for three
4 growing seasons to loblolly pine (Pinus taedd) at a field site in east Texas. After 3 years, the
5 trees were harvested and biomass measured. Neither EDU nor NaE prevented foliar O3 injury,
6 but EDU applications at 450 ppm resulted in increases in stem diameter and height and total
7 above-ground biomass. These measures of growth also tended to slightly increase with
8 applications of NaE, but the effects were nonsignificant.
9 Several recent studies used EDU in assessing the response of several plant species to O3 to
10 help validate the proposed critical level (AOT40 = 3000 ppb.h) for crop protection in Europe
11 (Ball et al., 1998; Ribas and Penuelas, 2000; Tonneijck and Van Dijk, 2002b; 2002a). EDU
12 appeared to provide protection from visible foliar injury, but the results regarding yield and
13 biomass reductions were mixed. In a 3-year study over 12 sites throughout Europe, Ball et al.
14 (1998) used the ratio of EDU-treated versus non-treated white clover (Trifolium repens)
15 biomass but did not find a significant relationship between biomass reductions and AOT40.
16 When other parameters, such as temperature and VPD were included in the model however,
17 they found a significant relationship (r2 = 0.79) between biomass reductions and AOT40, and
18 the greatest biomass reductions occurred in areas with the highest levels of industrialization
19 (in Germany).
20 Tonneijck and Van Dijk (2002b) assessed the relationship of visible injury of subterranean
21 clover (Trifolium subterraneum) to ambient O3 at four sites over three growing seasons in the
22 Netherlands, using EDU-treated and non-treated plants. Visible injury varied by site and year,
23 but was reduced to near zero by EDU treatment. However, no relationship indicative of a
24 protective effect of EDU with this plant species was observed for biomass. Tonneijck and
25 Van Dijk (2002a) also reported similar results with bean (Phaseolus vulgaris). Both EDU-
26 treated and non-treated plants were exposed to ambient O3 at three locations in Spain over one
27 growing season (Ribas and Penuelas, 2000). Reductions in yield and biomass were correlated
28 with O3 concentration and EDU provided some protective effect, although results varied by
29 location and with meteorological conditions.
30 The mechanisms by which protective chemicals, especially EDU, protect plants are poorly
31 understood. However, Wu and von Tiedemann (2002) reported that applications of two recently
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1 developed fungicides (azoxystrobin and epoxiconazole) provided protection of spring barley to
2 relatively high O3 exposures (150-250 ppb, 5 days, 7h/d) and resulted in increases in leaf soluble
3 protein content, as well as the activity of several antioxidative enzymes (e.g., superoxide
4 dismutase, catalase, ascorbate-peroxidase, and glutathione reductase). In addition, the increase
5 in these enzymes reduced superoxide levels in the leaves.
6 Before using these chemicals in a field setting, preliminary investigations should be done
7 to evaluate the methods and timing of application, as well as proper application rates, so as to
8 avoid any potential toxic effects (Manning, 2000; Manning and Krupa, 1992). Also, because
9 certain plant species may be nonresponsive, careful pre-screening of various species to be tested
10 is critical.
11 In summary, the use of protectants such as EDU (see Table 9-1) provides a ready means of
12 demonstrating the local occurrence of adverse effects of O3, both in terms of visible injury and
13 growth reduction. However, even if accompanied by some type of O3 and meteorological
14 monitoring, their use has yet to give rise to a methodology for quantitatively assessing exposure
15 response.
16
17 9.2.3.4 Biomonitoring
18 Bioindicators
19 The use of biological indicators to detect the presence of O3 injury to plants is a
20 longstanding and effective methodology (Chappelka and Samuelson, 1998; Manning and Krupa,
21 1992). A bioindicator can be defined as a vascular or non-vascular plant exhibiting a typical and
22 verifiable response when exposed to a plant stress such as an air pollutant (Manning et al., 2003).
23 To be considered a good indicator species, plants must:
24 (1) exhibit a distinct, verified response;
25 (2) have few or no confounding disease or pest problems; and
26 (3) exhibit genetic stability.
27 Such sensitive plants can be used to detect the presence of a specific air pollutant such as
28 O3 in the ambient air at a specific location or region and, as a result of the magnitude of their
29 response, provide unique information regarding specific ambient air quality. Bioindicators can
30 be either introduced sentinels, such as the widely used tobacco (Nicotiana tabacum) variety
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Table 9-1. Advantages and Disadvantages of Protective Chemicals Used in Assessment
of O3 Effects on Plants
Advantages
No chambers required. Plants exposed to ambient conditions of O3, light, temperature, etc.
Can conduct studies "in situ." Equipment needs are minimal. No "chamber effects"
A high degree of replication possible both within and among locations
Disadvantages
Exposure-response studies require inclusion of other methodologies (OTCs, etc.)
Need measurements of ambient O3 and other meteorological variables (temp, rainfall, etc)
Have to conduct preliminary toxicology studies to determine proper rate, timing etc.
Possible plant toxicity can result from repeated applications
Species response can vary; need to screen for proper species to use
Modified from Manning and Krupa (1992).
1 Bel W3, or detectors, which are sensitive native plant species. The approach is especially useful
2 in areas where O3 monitors are not operated (Manning et al., 2003). For example, in remote
3 wilderness areas where instrument monitoring is generally not available, the use of bioindicator
4 surveys in conjunction with the use of passive samplers (Krupa et al., 2001) is a particularly
5 useful methodology (Manning et al., 2003). However, the method requires expertise or training
6 in recognizing those signs and symptoms uniquely attributable to exposure to O3 as well as their
7 quantitative assessment.
8 Since the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996) many new
9 sensitive species have been identified from controlled exposure studies and verified in the field
10 (Flagler, 1998; Innes et al., 2001). In addition, several new uses of this methodology have been
11 demonstrated, including a national O3 bioindicator network, studies in wilderness areas, and
12 mature tree studies. Although it has been difficult to find robust relationships between the foliar
13 injury symptoms caused by O3 and effects on plant productivity or ecosystem function, visible
14 injury correlations with growth responses have been reported (Table 9-2; Chappelka and
15 Samuelson, 1998; Manning et al., 2003; Smith et al., 2003). One workshop on the utility of
16 bioindicators of air pollutants led to a useful series of peer-reviewed publications in
17 Environmental Pollution (Skelly, 2003).
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Table 9-2. Advantages and Disadvantages of Bioindicators Used to Study
O3 Plant Effects
Advantages
No chambers required. Plants exposed to ambient conditions of O3, light, temperature, etc.
Relatively inexpensive. Equipment needs are minimal. No "chamber effects"
A high degree of replication possible (sentinels) both within and among locations
Disadvantages
Individuals need to be trained and experienced in O3 symptom recognition
Need adequate numbers of plants (detectors) to ensure valid results
Need preliminary tests to insure a constant symptomotology of material used
Use more than one indicator species (detector) per area if possible
Quantify site characteristics (soils, light) that may influence symptom expression
Need measurements of ambient O3 (active or passive) and other meteorological variables
(temp, rainfall, etc)
Need to ensure cultural (sentinels) practices (soil, irrigation, fertilization, etc.) are similar
among sites
1 National network
2 The U.S. Forest Service in cooperation with other federal and state agencies developed a
3 network of O3 bioindicators to detect the presence of O3 in forested systems throughout the U.S.
4 (Smith et al., 2003). This ongoing program was initiated in 1994; and 33 states currently
5 participate. In a coordinated effort, a systematic grid system is used as the basis of plot
6 selection, and field crews are trained to evaluate O3 symptoms on sensitive plant species within
7 the plots (Coulston et al., 2003; Smith et al., 2003).
8 The network has provided evidence of O3 concentrations high enough to induce visible
9 symptoms on sensitive vegetation. From repeated observations and measurements made over a
10 number of years, specific patterns of areas experiencing visible O3 injury symptoms can be
11 identified. Coulston et al. (2003) used information gathered over a 6-year period (1994-1999)
12 from the network to identify several species that were sensitive to O3 over a regional scale
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1 included sweetgum (Liquidambar styraciflua), loblolly pine (Pinus Taeda), and black cherry
2 (Prunus serotind).
3
4 Wilderness areas
5 The use of bioindicator species as detectors has proven to be an effective technique for
6 deriving a relative estimate of O3 injury in wilderness areas in both the US and Europe
7 (Chappelka et al., 1997; 2003; Manning et al., 2002). However, to be truly effective, these
8 regional and national bioindicator studies need the inclusion of air quality data and related
9 growth studies to determine effects on productivity and ecosystem function (Bytnerowicz et al.,
10 2002; Manning et al., 2003; Smith et al., 2003).
11 Chappelka et al. (1997; 2003) conducted surveys of foliar injury on several native plant
12 species throughout the Great Smoky Mountains National Park (GRSM), including black cherry
13 (Prunus serotind), tall milkweed (Ascelpias exaltatd), cutleaf coneflower (Rudbeckia laciniatd)
14 and crownbeard (Verbesina occidentalis). Visible foliar symptoms were prevalent throughout
15 the Park, indicating that injury-producing O3 levels were widespread in GRSM.
16 Manning et al. (2002) recently summarized a multi-year (1993-2000) bioindicator project
17 in the Carpathian Mountain range in eastern Europe. They evaluated numerous trees, shrubs,
18 forbs, and vines for possible symptoms of O3 injury. Observations were made at plots located in
19 the vicinity of either active or passive O3 monitors (Bytnerowicz et al., 2002). Approximately 30
20 species of native plants detectors were identified as possible bioindicators, the majority of which
21 (21) were shrubs (Manning et al., 2002). Based on these observations, it was concluded that O3
22 concentrations were sufficiently high to impact ecosystems in the region. Similar investigations
23 regarding the sensitivity of native species have been conducted in Switzerland (Novak et al.,
24 2003) and Spain (Orendovici et al., 2003).
25
26 Mature tree detectors
27 Many studies have reported visible injury of mature coniferous trees caused by O3,
28 primarily in the western United States (Arbaugh et al., 1998) and, to a lesser extent, to mature
29 deciduous trees in eastern North America. In an effort to determine the extent and magnitude of
30 visible injury in mature tree canopies, Hildebrand et al. (1996) and Chappelka et al. (1999b)
31 conducted independent studies in the GRSM and the Shenandoah National Park (SHEN). The
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1 species examined were sassafras (Sassafras albiduni), black cherry, and yellow-poplar
2 (Liriodendron tulipifera L.) in GRSM and white ash (Fraxinus americanaZ.), black cherry, and
3 yellow-poplar in SHEN. Protocols were similar at both parks, and trees were located near O3
4 monitors at three different areas in each park. Results from both studies indicated that symptoms
5 of O3 injury were present in the trees and correlated with the amount of injury both spatially and
6 temporally. Ozone injury tended to be most severe at the highest elevation, except with yellow-
7 poplar.
8 Hildebrand et al. (1996) observed significant O3 exposure-plant response relationships with
9 black cherry. The best relationships were found between foliar injury and the SUM06 and W126
10 exposure indices, indicating that higher O3 concentrations were important in eliciting a response
11 in black cherry. No O3 exposure-plant response relationships were found with any species tested
12 in GRSM (Chappelka et al., 1999b); but, when the data were combined for both parks, a
13 significant correlation r = 0.72) with black cherry was found for both SUM06 and W126, and
14 injury was the greatest (r = 0.87) at the higher elevations (Chappelka et al., 1999a).
15 Based on a study in which visible symptoms of O3 injury were characterized for large,
16 mature yellow-poplar and black cherry trees in GRSM (Chappelka et al., 1999a), Somers et al.
17 (1998) compared radial growth differences among trees classified as sensitive or non-sensitive
18 based on the severity of visible foliar injury observed over a 3-year period (1991-1993).
19 Significantly more radial growth was observed over both a 5- and a 10-year period for the non-
20 sensitive compared to the sensitive trees. No significant relationship was found for black cherry
21 tree growth.
22 Vollenweider et al. (2003), using data collected from continuous forest inventory (CFI)
23 plots across Massachusetts, compared growth rates among either symptomatic or asymptomatic
24 mature black cherry trees. Of the 120 trees sampled in 1996, 47% exhibited visible foliar injury.
25 Using CFI data, growth rates were compared over a 31-year period. The growth rates for
26 symptomatic trees were reduced by 28% compared with the asymptomatic trees.
27 Because these studies (Somers et al., 1998; Vollenweider et al., 2003) were not controlled
28 studies and used a small sample of trees, they cannot validly be used to characterize cause and
29 effects related to the visible symptoms and radial growth they describe. However, the results
30 indicate the possibility that O3 is correlated with growth losses in some sensitive genotypes,
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1 illustrating the potential usefulness of this visible O3 injury methodology in assessing effects on
2 the growth rates of mature deciduous trees.
3
4 Cultivar Comparisons
5 The idea of using cultivars or isogenic lines of crop species that differed in O3 sensitivity as
6 sentinels to determine the ambient effects of O3 in the field was presented in the 1996 O3 AQCD
7 (U.S. Environmental Protection Agency, 1996). The rationale was that comparing the ratio of
8 injury scores or some measure of growth between two different cultivars varying in O3
9 sensitivity should be indicative of the relative amount of ambient stress to plants at a given
10 location. A sensitive:resistant ratio close to unity would indicate relatively low O3
11 concentrations and a low ratio higher O3 levels. Results from locations differing in O3
12 concentrations could be evaluated to develop exposure-response models. The original protocol
13 was derived using two isogenic lines of white clover (Trifolium repens) differing in O3
14 sensitivity (Heagle et al., 1994b; 1995).
15 This white clover model system has been used in several multi-location studies in the U.S.
16 (Heagle and Stefanski, 2000) and Europe (Ball et al., 2000; Bermejo et al., 2002; Mills et al.,
17 2000). Heagle and Stefanski (2000) compared results from eight sites over a 2-year period with
18 various exposure indices (SUMOO, SUM06, W126 and others) to determine a best-fit regression.
19 They found that most of the indices preformed similarly. The highest r2 values (0.87-0.93) were
20 obtained using only the later harvests and a 6 h d"1 index (1000-1600 h). Similar multiple-
21 comparison studies conducted in Europe using the AOT40 index (Ball et al., 2000; Mills et al.,
22 2000) yielded poorer r2 values. Factors such as air temperature, NOX (high levels at some sites)
23 and lower O3 concentrations in Europe were suggested to account in part for the differences
24 between U.S. and Europe study results. Bermejo et al. (2002), in a study in Spain, improved the
25 model by comparing the biomass ratio of these white clover isolines to measures of O3 uptake
26 (flux) rather than an exposure index (AOT40). Together, these studies indicate that systems
27 such as the white clover model can help reveal O3 exposure-response relationships that can
28 provide valuable information regarding ambient O3 conditions in a given location. Table 9-3
29 lists the advantages and disadvantages of the use of cultivar comparisons in assessing O3 effects
30 of pi ants.
31
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Table 9-3. Advantages and Disadvantages of Cultivar Comparisons Used in
Assessment of O3 Effects on Plants
Advantages
No chambers required. Plants exposed to ambient conditions of O3, light, temperature, etc.
Relatively inexpensive. Equipment needs are minimal. No "chamber effects"
A high degree of replication possible both within and among locations
Can conduct studies "in situ"
Disadvantages
Preliminary tests to insure sensitivity and growth patterns of genotypes used are consistent
are needed
Need measurements of ambient O3 and other meteorological variables (temp, rainfall, etc)
Have to ensure cultural practices (soil, irrigation, fertilization, etc.) are similar among sites
Closely monitor plants for presence of other factors that may cause a misinterpretation of results
1 Dendrochronological techniques
2 It has been difficult to determine whether O3 significantly affects tree growth and
3 productivity in the field, because O3 concentrations are omnipresent and tree response to this
4 pollutant is altered by many factors. The use of dendrochronological techniques to answer
5 questions regarding ambient O3 effects on forest growth and ecosystem function has recently
6 emerged as a very useful biomonitoring methodology (Cook, 1990; McLaughlin et al., 2002).
7 The technique is useful when either instrument or passive O3 monitoring methods are used to
8 determine ambient O3 conditions.
9 Initial experiments were primarily correlative in nature and attempted to relate symptoms
10 of visible injury with growth losses as revealed by tree ring analysis (Arbaugh et al., 1998;
11 Benoit et al., 1983; Peterson et al., 1995; Somers et al., 1998; Swank and Vose, 1990). These
12 studies evaluated radial growth patterns determined by cores removed from trees in the presence
13 or absence of overt O3 injury symptoms.
14 The method has also been adapted to better understand forest ecosystem function
15 (McLaughlin and Downing, 1996; McLaughlin and Downing, 1995; Bartholomay et al., 1997;
16 McLaughlin et al., 2003). The response of mature loblolly pine (Pinus Taedd) growing in
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1 eastern Tennessee to ambient O3 and moisture stress was evaluated by McLaughlin and Downing
2 (1996, 1995) over a 6-year period (1988 to 1983). They made radial growth measurements
3 from 12 to 37 times per year using dendrometer bands and determined relationships between O3,
4 moisture stress, and radial growth. Exposures to O3 concentrations > 0.04 |_il L/1 with high
5 temperatures and low soil moisture resulted in short-term depression in radial growth.
6 Reductions in growth were estimated to vary from 0 to 15% per year and averaged
7 approximately 5% per year.
8 Bartholomay et al. (1997) examined white pine (Pinus strobus) radial growth in eight
9 stands throughout Acadia National Park, Maine over a 10-year period from 1983 to 1992. They
10 related growth rates to several factors, including O3 concentration. Ozone levels were negatively
11 correlated with radial growth in seven of the eight stands. Site characteristics were important in
12 the relationship: stands growing on shallow, poorly drained soils were most sensitive to O3 in
13 the late portion of the growing season, possibly due to premature senescence of foliage.
14 However, litterfall measurements were not reported. Trees growing on better sites were more
15 sensitive to O3 during the entire growing season, indicating the possibility of high O3 uptake
16 rates throughout the growing season. Although these field studies (Bartholomay et al., 1997;
17 McLaughlin and Downing, 1996; McLaughlin and Downing, 1995) did not compare the firm O3
18 effects on the two pine species, they indicate that potential interactions exist among O3 and other
19 climatic and edaphic factors, such as temperature and soil moisture.
20 Using both automated and manual dendrometer bands, McLaughlin et al. (2002) examined
21 the growth response of yellow poplar (Liriodendron tulipifera) trees recently released from
22 competition. In addition to measuring growth, sap flow measurements were conducted and soil
23 moisture was measured in the vicinity of the trees. However, O3 concentrations were low and
24 there were no negative O3 effects observed in this 1-year study. Advantages and disadvantages
25 of dendrochronology techniques for evaluating whole-tree physiological responses for individual
26 trees and forest stands are listed in Table 9-4.
27 The use and evolution of various dendrochronological methods in the field of air pollution
28 effects research is reviewed in detail by McLaughlin et al. (2002). Automated dendrometer
29 bands provide a powerful tool for measuring radial growth responses of trees on an hourly or
30 daily basis. Diurnal patterns of growth can be related to water use and O3 concentrations using
31
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Table 9-4. Advantages and Disadvantages of Various Dendrochronological Techniques
Used in Assessment of O3 Effects on Plants
Advantages
Provide information regarding growth effects under ambient conditions
Good historical information regarding O3 effects
Can provide data on daily, and seasonal growth and O3 patterns and relate these to
physiological function
Provide information on forest function related to ambient O3 concentrations
Can link data with process level growth models
Disadvantages
Results are generally correlative in nature with no true control
Need background O3 and meteorological data (historical records)
Need to account for other factors such as competition, in analyzing data
Individuals need to be trained in counting growth rings
Replication can be difficult (expensive and technological limitations)
Complicated statistical analyses are sometimes required
Can be expensive, especially if using automated growth (dendrometer) bands
1 time-series analyses. The major drawbacks of the method are that it is expensive and time
2 consuming.
3
4 9.2.3.5 Calibrated Passive Monitors
5 Many studies have used passive monitors in the mapping of ambient O3 concentrations,
6 especially in remote areas (Cox and Malcolm, 1999; Grosjean et al., 1995; Krupa et al., 2001).
7 Since they are cumulative recording devices, they cannot reveal short-term variations in O3
8 concentration but only the total exposure over a given interval, usually 7 days. Thus, they
9 produce a measurement that resembles the instrumentally derived exposure SUMOO index.
10 Runeckles and Bowen (2000) used the ZAPS system described in Section 9.2.2.3 to subject
11 both crops and passive monitors (Williams, 1994) to a range of exposures. Passive monitors
12 were also exposed at 16 agricultural field sites along a transect through the Fraser Valley, British
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1 Columbia, Canada. Most field sites were downwind of the Greater Vancouver metropolitan
2 area. All passive monitors were replaced at weekly intervals and the data from those in the
3 ZAPS plots were "calibrated" to crop responses by means of Weibull exposure-response
4 functions. Since the meteorological conditions throughout the valley were reasonably consistent
5 from site to site, the use of these functions with data from the network passive monitors as inputs
6 permitted the estimation of crop losses at the network sites. The overall method was thus a
7 hybrid of several methodologies.
8 Although based on a single study, the use of passive monitors has potential for assessing
9 crop losses at sites removed from locations with known ambient O3 concentrations. Provided
10 that the network and calibration sites have similar meteorological conditions, the method yields
11 crop loss estimates that are responses to local ambient O3 levels as influenced by local
12 meteorological conditions
13
14 9.2.4 Numerical/Statistical Methodologies
15 Proper experimental design strategies including replication, randomization, and
16 experimental protocols are paramount in O3-effects research. These have been discussed in detail
17 in previous O3 AQCDs (U.S. Environmental Protection Agency, 1986, 1996), as have the
18 different statistical analytical procedures used to determine the probable significance of results.
19 However, new investigative approaches have demanded the adoption of new analytical methods.
20 For example, the use of dendrochronological techniques has led to the use of time-series analysis
21 (McLaughlin et al., 2003) and linear aggregate models (Cook, 1990), as reviewed by
22 McLaughlin et al. (2002).
23 In spite of the rigors of the analyses, many differences occur in the published literature for
24 almost any plant response to O3 stress. Differences inevitably result from different researchers
25 studying different locations, using different experimental methodologies and genetically
26 different plant material even when using a common species. The techniques of meta-analysis
27 can be used to consolidate and extract a summary of significant responses from a selection of
28 such data.
29 Despite the differences in responses in the 53 primary studies used, a recent meta-analysis
30 by Morgan et al. (2003) of the effects of O3 on photosynthesis, growth, and yield of soybean
31 (Glycine max) showed "overwhelming evidence for a significant decrease in photosynthesis, dry
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1 matter production and yield ... across all the reported studies on effects of chronic O3 treatment."
2 The meta-analysis defined O3 stress exposure to -70 ppb for at least 7 days and found average
3 shoot biomass and seed yield decreases of 34% and 24%, respectively. Furthermore, although
4 other stress factors such as drought and UV-B did not affect the O3 responses, elevated CO2 was
5 found to significantly decrease O3-induced losses.
6 The meta-analysis method clearly has the potential to consolidate and refine the
7 quantitative exposure-response models for many species. The majority of the reported growth
8 and physiological responses related to O3 stress are for individual plants, primarily in various
9 types of exposure chambers. It is difficult to extrapolate these responses to stand/community,
10 ecosystem, or region-wide assessments, particularly in view of the importance of the significant
11 interactions that may occur between plant responses O3 and other environmental stresses. Along
12 with the shift in effects research to a more ecological approach, these concerns necessitate a
13 move from simple regression analysis to more complex mathematical approaches to handle a
14 wider array of independent input variables than O3 exposure alone. Other independent input
15 variables that must be accounted for include air and soil temperatures, soil moisture, relative
16 humidity, wind speed, and, particularly in the case of natural systems, biotic factors such as pests
17 and pathogens, plant density/spacing, and measures of plant competition.
18 Artificial neural network (ANN) methodology was used by Balls et al. (1995) for
19 "unraveling the complex interactions between microclimate, ozone dose, and ozone injury in
20 clover" and in the study with the protectant chemical EDU, discussed in Section 9.2.3.3 (Ball
21 et al., 1998). The multi-factor model for predicting the effects of ambient O3 on clover
22 developed by Mills et al. (2000) utilized both ANN and multiple linear regression methods.
23 Models incorporating ANNs are of the "regression" type (Luxmoore, 1988) in contrast to
24 "mechanistic" or "phenomenological" models which have wider applicability. Process-level
25 models of either type have been developed at the organelle, individual plant (Constable and
26 Taylor, Jr., 1997; Weinstein et al., 1998), canopy (Amthor et al., 1994), and stand level (Ollinger
27 et al., 1997; Weinstein et al., 2001) and provide estimates of the rate of change of response
28 variables as affected by O3 over time. However, as pointed out in the 1996 Criteria Document
29 (U.S. Environmental Protection Agency, 1996), mechanistic process models lack the precision of
30 regression models as well as their ability to estimate the likelihoods of responses. In their
31 extensive reviews, Kickert and Krupa (1991) and Kickert et al. (1999) summarized the
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1 advantages and shortcomings of many different models and make the important point that most
2 of the models that have been described provide consequence assessments that quantify the
3 magnitudes of effects, but not risk assessments that quantify the likelihoods of such effects.
4 Descriptions of several specific models are provided in other sections of this criteria document,
5 and advantages and disadvantages of modeling techniques used in assessing O3 effects on plants
6 are summarized in Table 9-5.
7
Table 9-5. Advantages and Disadvantages of Modeling Techniques Used in Assessment
of Oa Effects on Plants
Advantages
• Provides an understanding of cause-effect relationships over time
Disadvantages
• Have to make assumptions based on scarcity of data
• Most models are very complex and difficult to understand
• Need to be evaluated for predictive validity
1 9.2.5 Improved Methods for Defining Exposure
2 Ambient air quality is defined in terms of the measured O3 concentrations in the air at some
3 standard elevation above ground level. Compilations of such concentration data have long been
4 used as surrogates of the exposures to which plants are subjected. However, as long ago as
5 1965, field research provided evidence that plant response was a function, not of ambient O3
6 concentration/>er se, but of the estimated flux of O3 to the plant canopy (Mukammal, 1965).
7 Subsequently, Runeckles (1974) introduced the term "effective dose" to define that part of the
8 ambient exposure that was taken up by a plant. Fowler and Cape (1982) later referred it as
9 "pollutant applied dose" (PAD), defined as the product of concentration, time and stomatal
10 (or canopy) conductance, with units g m"2. Such estimates of O3 uptake or flux provide a more
11 biologically relevant description of exposure than the simple product of concentration and time
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1 alone, and formed the basis of Reich's 1983 "unifying theory" of plant response to O3 (Reich,
2 1983).
3 However, it was not until the early 1990s that the inherent advantages of using O3 flux
4 rather than O3 concentration as a basis for determining response effects began to be widely
5 accepted, as demonstrated by the subsequent increase in publications involving flux
6 measurements and modeling (e.g., Fuhrer et al., 1997; Griinhage and Jager, 2003; Griinhage
7 et al., 1993, 1997; Massman et al., 2000; Musselman and Massman, 1999; Pleijel, 1998). A key
8 requirement for flux determination is the measurement of stomatal or canopy conductances,
9 using established porometer/cuvette techniques or eddy correlation methods. The usefulness and
10 relevance of flux as a measure of exposure are discussed in detail in Section 9.5.
11 Efforts to develop models of O3 deposition and stomatal uptake are currently under way
12 with a view to providing improved assessments of the risks to vegetation across Europe
13 (Emberson et al., 2000; Simpson et al., 2001; Simpson et al., 2003).
14
15
16 9.3 SPECIES RESPONSE/MODE-OF-ACTION
17 9.3.1 Introduction
18 The evaluation of O3 risk to vegetation requires fundamental understanding of both the
19 functioning of the vegetation and how external environmental influences can alter that function.
20 For biological organisms subjected to atmospheric O3, those alterations can be complex and
21 multiple. In addition, biological organisms have plasticity to external interactions due to their
22 complex internal, self-correcting systems, making the task of identifying their "correct"
23 functioning difficult. This section emphasizes reactions of O3 with the cell and tissue rather than
24 the whole plant to describe the fundamental mechanisms known to govern the response of the
25 plant to O3 exposure.
26 The many regulatory systems contained in leaves change both as a function of leaf
27 development and in response to various environmental stresses. Leaves function as the major
28 regulators of anatomical and morphological development of the shoot and control the allocation
29 of carbohydrates to the whole plant (Dickson and Isebrands, 1991). This section discusses the
30 movement of O3 into plant leaves and their biochemical and physiological responses to O3.
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1 The 1996 criteria document (U.S. Environmental Protection Agency, 1996) assessed the
2 information available at that time concerning the biochemical and physiological responses to the
3 movement of O3 into plant leaves. This information continues to be valid. Ozone uptake in a
4 plant canopy is a complex process involving adsorption to surfaces (leaves, stems, and soil) and
5 absorption into leaves (Figure 9-1). However, the initial biochemical changes that result within
6 leaf cells after the entry of O3 and how these changes interact to produce plant responses remain
7 unclear. The response of vascular plants to O3 may be viewed as the culmination of a sequence
8 of physical, biochemical, and physiological events. Only the O3 that diffuses into a plant
9 through the stomata (which exert some control on O3 uptake) to the active sites within a leaf
10 impairs plant processes or performance. An effect will occur only if sufficient amounts of O3
11 reach sensitive cellular sites that are subject to the various physiological and biochemical
12 controls within the leaf cells. Ozone injury will not occur if (1) the rate and amount of O3 uptake
13 is small enough for the plant to detoxify or metabolize O3 or its metabolites, or if (2) the plant is
14 able to repair or compensate for the O3 impacts (Tingey and Taylor, 1982; U.S. Environmental
15 Protection Agency, 1996). Therefore, a precondition for O3 to affect plant function is that it
16 must enter the stomata and be absorbed into the water lining the mesophyll cell walls. The
17 response of each plant is determined by the amount of O3 entering the leaves, which varies from
18 leaf to leaf.
19 Some potentially significant processes have been investigated since the 1996 criteria
20 document, especially detoxification and compensatory processes. The role of detoxification in
21 providing a level of resistance to O3 has been investigated; however, it is still not clear as to what
22 extent detoxification can protect against O3 injury. Data are needed especially on the potential
23 rates of antioxidant production and on the subcellular localization of the antioxidants. Potential
24 rates of antioxidant production are needed to assess whether they are sufficient to detoxify the O3
25 as it enters the cell. The subcellular location(s) are needed to assess whether the antioxidants are
26 in cell wall or plasmalemma locations that permit contact with the O3 before it has a chance to
27 damage subcellular systems. Various forms of compensation, especially the stimulation of new
28 leaf production and of higher photosynthetic performance of new leaves, have been reported.
29 Although these processes divert resources away from other sinks, these forms of compensation
30 may counteract the reduction in canopy carbon fixation caused by O3. The quantitative
31 importance of these processes requires investigation.
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Atmospheric Processe
i
[Ozone in the | f Canopy |
5 "" 1 Atmosphere 1 1 Boundary Layer 1
r
Canopy Processes
i
r
Leaf Processes/
Ozone Uptake S
^
r
Leaf Processes/
Mode of Action
I
Plant Response
\
Ecosystem Responses
•*^^
"** *«N
k
\
\
\
\
\
\
\
\
\
\
\
~D>
^
K Canopy
Conductance
r :-' -
/^Ozone in N
Vegetation
\^ Canopy J
[>
T
^,
J y
•• • ji p '- ' •-
*, \&f -,',•:'---
~>^!
f
f Leaf Boundary Layer J
S
^^ Leaf
^\_ Conductance
V7
r C Rtn
(f ^ 1 Cond
Ozone Absorption V
/ r-
?^
^^L
matal |
uctance I
1 into Foliage 1
Figure 9-1. Ozone uptake from the atmosphere. Ozone moves from the atmosphere
above the canopy boundary layer into the canopy primarily by turbulent
air flow. Canopy conductance, controlled by the complexity of the canopy
architecture, is a measure of the ease with which gases move into the canopy.
Within the canopy, O3 is adsorbed onto surfaces as well as being absorbed
into the foliage. Foliage absorption is controlled by two conductances, leaf
boundary layer and stomatal, which together determine leaf conductance.
The solid black arrows denote O3 flow; dotted arrows indicate processes
affecting uptake or response to O3. Boxes at the left with double borders
are those processes described in the figure.
1 As a result of the research since the 1996 criteria document (U.S. Environmental Protection
2 Agency, 1996), the way in which O3 exposure reduces photosynthesis, especially its effects on
3 the central carboxylating enzyme, Rubisco (ribulose-l,6-P2-carboxylase/oxygenasel), is better
4 understood. The rate of leaf senescence has been shown to increase as a function of increasing
5 O3 exposure. The mechanism of the increased senescence is not known, and, hence, it deserves
6 further study.
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1 Finally, the role that changes in allocation of resources play in plant response to O3 is now
2 better understood. Most studies have shown that O3 decreases allocation of photosynthate to
3 roots. In some cases, allocation to leaf production has increased. Whether these changes are
4 driven entirely by changes in carbohydrate availability or are controlled by other factors (e.g.,
5 hormones) is not known. Physiological effects within the leaves inhibit photosynthesis, alter the
6 assimilation of photosynthate and shift its allocation patterns, and can lead to reduced biomass
7 production, growth, and yield (U.S. Environmental Protection Agency, 1986, 1996).
8 The major problem facing researchers trying to predict long-term O3 effects on plants is
9 determining how plants integrate the responses to O3 exposures into the overall naturally
10 occurring responses to environmental stressors. Little is now known about how plant responses
11 to O3 exposures change with increasing age and size, but this information is crucial to predicting
12 the long-term consequence of O3 exposure in forested ecosystems.
13 This section focuses on reactions of O3 within cells and cellular tissue, in order to explain
14 known mechanisms that govern plant responses. The processes that occur at cell and tissue
15 levels within the leaf will be divided into several steps beginning with O3 uptake and its initial
16 chemical transformations into a series of currently unknown, but suspected toxic, chemicals
17 (Figure 9-2). The discussion will then focus sequentially upon various cell regions, their general
18 physiology, and the changes that may occur within a plant after O3 exposure. This is important
19 because the varying responses of the different plant species in a community ultimately lead to
20 an ecosystem response. Finally, a general summary is presented that discusses the known or
21 suspected changes that occur within the whole plant.
22
23 9.3.2 Mechanisms of Ozone-Induced Plant Alterations
24 Plant adaptations for surviving O3 stress include exclusion or tolerance of it or its products
25 (Levitt, 1972; Tingey and Taylor, 1982). Ozone may be excluded from tissues or cells via
26 stomatal closure, by extracellular oxidants, or by membrane impermeability to O3 or its products.
27 Past investigations of O3 injury have indicated that physiological and metabolic changes occur
28 (see Heath, 1998; Reddy et al., 1993; Harris and Bailey-Serres, 1994). Many of these changes
29 are likely initiated via gene expression. During the last decade, our understanding of the cellular
30 processes within plants has increased. Although the fundamental hypotheses concerning
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Atmospheric Processes
I Canopy Processes |
X
I Leaf Processes | J
x
I Plant Response Processes |
1
| Community Processes
4
I Ecosystem Processes
Plasma Membrane
Second Sites:
Signaling
Second Sites:
Metabolism
General Physiology:
Cell Metabolism
General Physiology:
Whole Plant
Figure 9-2. Absorption and transformation of O3 within the leaf. The varied processes are
broken down in to smaller mechanistic steps that lead from uptake of
atmospheric O3 into the alterations which may occur within the individual
plant. Each plant responds to the O3 level and therefore interacts with the
total ecological setting to generate an ecosystem response due to the O3.
1
2
3
4
5
O3-induced changes in physiology have not changed, a more complete development of the
theories is now nearing possibility.
9.3.2.1 Changes in Metabolic Processes: Current Theories
The current hypotheses regarding the biochemical response to O3 fumigation revolve about
injury and its prevention. These are well discussed by Pell et al. (1997) and are listed below in
no order of importance. Although they are listed separately, some may be interlinked and related
to each other.
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1 (1) Membrane Dysfunction. The membrane is altered by O3, principally via protein
changes not involving the lipid portions of the membrane (except at extremely high
levels of O3). These alterations involve increased permeability with perhaps lessened
selectivity, declines in active transport, and changes in the trigger mechanisms of
signal transduction pathways such that the signals are no longer suitable for the state
of the cell. The cellular pools and transport systems of Ca2+/K+/H+ are the primary
suspects.
2 (2) Antioxidant Protectants. Varied antioxidants (both as metabolites and enzyme
systems) can eliminate the oxidant or its products, if present at time of fumigation
and in the sufficient abundance. However, oxidant entry that occurs rapidly can
overwhelm the antioxidant response.
3 (3) Stress Ethylene Interactions. Visible injury is caused by the interaction of O3 with
stress-induced ethylene, either by direct chemical transformation to a toxic product or
by alteration of the biochemical relations at the ethylene binding site.
4 (4) Impairment of Photosynthesis. A product of O3 (and less probably, O3 itself) enters
the cell, causing a decline in the mRNA for Rubisco (especially the message RNA
species ofrbcS and rbcL) such that Rubisco levels slowly decline within the
chloroplast, leading to a lowered rate of CO2 fixation and productivity. This process
is very similar to early senescence and may be linked to general senescence.
Alternatively, a false signal is generated at the cell membrane which lowers the
transcription of DNA to mRNA. Ozone alters the normal ionic and water relations of
guard cells and subsidiary cells, causing the stomata to close and limit CO2 fixation.
In any case, the response of the stomata to the current environment does not promote
efficient photosynthesis.
5 (5) Translocation Disruption. One of the biochemical systems most sensitive to O3
exposure is the translocation of sugars, such that even a mild exposure inhibits the
translocation of carbohydrate (Grantz and Farrar, 1999, 2000).
6 (6) General Impairment/Disruption of Varied Pathways of Metabolism. This is the
oldest and most vague concept of how O3 alters metabolism. It is based upon early
work in which what enzymes and metabolites could be assayed were. Thus, these
results were based upon what could be done, rather than a coherent hypothesis. The
best examples are listed in Dugger and Ting (1970).
7 The latter two theories can be restated as a loss of productivity with three possible
8 somewhat-independent causes — (a) a reduced production of the basic building blocks of growth
9 and, hence, a slowing of growth in at least one organ; (b) a reduced ability to reproduce, leading
10 to a decreased production of viable seeds or of fruits and nuts; and (c) a decreased ability to
11 mount a defense against pathogens or insects, leading to weaker plants, which are more liable to
12 be overcome by other stresses. It is important to separate out effects that may be detrimental or
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1 disfiguring, such as the production of visible injury, but which have not been shown to lead
2 directly to a loss of productivity due to possible compensation by the remaining tissue. The loss
3 of some localized regions of photosynthesis tissue has not been shown to lead directly to a loss
4 of productivity.
5
6 9.3.2.2 Modifications of Plant Physiological Processes
7 The discussion that follows will focus on physiological processes; the species used to
8 develop an understanding of these processes are relatively unimportant. The study of any plant
9 that can help increase the biological understanding of the response of plants to O3 exposures is of
10 use, regardless of how sensitive it is to O3. Therefore, Arabidopsis, whose physiology and
11 genome continue to be studied and described by a large number of scientists is an appropriate
12 plant for studying O3 injury. Though the responses of mature trees and understory plants are
13 critical to understanding plant interactions at an ecosystem level, the time required for trees to
14 reach maturity makes using them to study biological mechanisms a poor choice.
15 The high levels of O3 used for some investigations do not automatically invalidate the
16 results obtained in those studies. Typically when a new hypothesis is being investigated,
17 extreme levels of the toxicant are used to clearly determine its effects. The older studies that
18 used concentrations as high 1 ppm, an extreme level, helped to define current studies. Later
19 experiments have used concentrations nearer ambient levels. Many of the current studies on
20 physiology use exposures between 0.15 and 0.25 ppm, which though higher than ambient levels
21 in some areas of the country, bypass confounding changes but allow for rapid experiments.
22 Three forms of air pollutant-induced injury patterns currently are known to exist: (a) acute
23 stress, generated by high atmospheric concentrations of pollutants for short periods of time;
24 (b) chronic stress, generated by lower concentrations of pollutants for long periods of time; and
25 (c) accelerated senescence, generated by very low concentrations of pollutants for very long
26 periods of time. At higher levels, distinct visible injury generally occurs due to cellular and
27 tissue death of regions of leaf mesophyll cells. This leads to a decline in the total area of
28 metabolically active tissue, with consequent loss of membrane integrity, loss of metabolites into
29 the extracellular tissue space, and formation of oxidative products. When no visible injury is
30 observed, lowered rates of photosynthesis or productivity are often used to index injury. Under
31 these conditions, metabolism is altered and the pool sizes of many metabolites are changed.
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1 More importantly, the altered biochemical states within the tissue lead to the inability of the
2 plant to respond properly to existing environmental conditions and to other stressors (Manning
3 and Keane, 1988; Heath, 1988; 1994b; Koziol and Whatley, 1984; Schulte-Hosted et al., 1988).
4
5 9.3.3 Ozone Uptake by Leaves
6 Plants respond to O3 similarly to other stressors on several levels: exclusion, tolerance, and
7 repair (Levitt, 1972; Tingey and Taylor, 1982). The response mechanism depends upon the O3
8 concentration, environmental conditions, and the developmental and metabolic state of the plant
9 (Guzy and Heath, 1993). These responses are detrimental to plant productivity because they cost
10 the plant metabolic resources. In some cases, the stomata close under the O3 exposure,
11 excluding the pollutant from the leaf interior and preventing injury. However, if this happens
12 too often, CO2 fixation is also inhibited and plant productivity suffers.
13 Atmospheric O3 does not cause injury, but rather it is the O3 that enters the plant that
14 causes an effect (Tingey and Taylor, 1982). Three well-defined, sequential processes control the
15 movement of O3 from the atmosphere into the sites of action within the leaf and must occur to
16 trigger O3 stress (Heath, 1980). The processes are: (1) entry of O3 into the leaf, (2) reactions of
17 O3 and its possible reaction product(s) in the water phase at cell surfaces, and (3) movement of
18 O3 reaction product(s) into the cell with enzymatic or chemical transformation of those products
19 in the cell.
20
21 Process 1. Entry ofO3 into the leaf. Often incorrectly, the external concentration of O3 is
22 used to give an indication of "dose" (Heath, 1994a). Ozone-induced changes on the plants
23 cuticle are minimal, and O3 does not penetrate the cuticle (Kerstiens and Lendzian, 1989) to
24 cause an effect. As O3 has no easily measured isotope, virtually no measurements have been
25 done on an actual dose of O3, i.e., the amount of O3 which reacts with individual biochemicals in
26 the leaf. Yet the measurement of dose will be the amount of O3 expected to penetrate into the
27 tissue through the stomata. Dose is expressed as a rate of delivery to a surface area (mol/m2 s"1).
28 Whether dose or total accumulation (mole/m2, rate integrated over exposure time) is most critical
29 for the development of injury remains a major question.
30 Ozone uptake includes gaseous diffusion through the leaf boundary layer and stomata into
31 the substomatal cavity (Figure 9-3). Although the movement of pollutants through a boundary
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Bulk Air Z^>
Boundary L-hv
Layer \~^
Stomatal 1— 1\
Aperture V~r
Sub-
stomatal
Cavity
Wind Speed Leaf
Canopy Temperature
Structure
Figure 9-3. The uptake of O3 into the leaf. Each of the individual concentration
layers of O3 represents a different process of movement and of plant/
microenvironmental interaction. This figure leads into Table 1, in which
the amounts of O3 along the pathway are calculated.
1 layer into the stomata region is known to be important, and even rate limiting in many cases of
2 low wind velocity, its description has been defined from aeronautical concepts and usually
3 relates to smooth surfaces that are not typical of leaf-surf ace morphology; however, it is nearly
4 the only treatment available (Gates, 1968). Once through the boundary layer, the gas must enter
5 the leaf through the stomata. The entry of gases into a leaf is dependent upon the physical and
6 chemical processes of gas phase and surfaces and is a well defined path that approximately
7 follows a linear flux law of:
8
j = g(C0-Ci) (9-1)
9
10 where the flux,y, into the internal space of a leaf is related to the conductance, g, through the
11 boundary layer and stomata and the gradient of concentration of gas from the outside, C0,
12 inwards, C;. This formulation has been used for years for both water and CO2 (Figure 9-4), and
13 for regions of varied CO2 concentration that correspond to C0 (CO2 of the atmospheric air, below
14 the leaf proper) and Q (CO2 near the leafs spongy mesophyll cells) (Farquar and Sharkey, 1982;
15 Ball, 1987).
16 In the past, the internal concentration of O3 has been assumed to be zero (Laisk et al.,
17 1989), due to early studies that found that virtually no O3 could pass through a leaf. That was
18 expected because O3 is extremely reactive with cellular biochemicals. If the assumption that the
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Light
Cuticle
Epidermis r
Pallisade
Mesophyll
Spongy
Mesophyll
Epidermis
Cuticle
nnnnnn
Vascular
System
C0=[C02]-
Figure 9-4. The microarchitecture of a dicot leaf. While details among species vary,
the general overview remains the same. Light that drives photosynthesis
generally falls upon the upper (adaxial) leaf surface. Carbon dioxide and
O3 enters through the stomata on the lower (abaxial) leaf surface, while
water vapor exits through the stomata (transpiration).
1 internal concentration zero is correct, then the effective delivery rate for O3 is given as g x C0,
2 with stomatal conductance being the major regulatory control (Taylor, Jr. et al., 1982; Amiro
3 et al., 1984). However, a recent study by Moldau and Bichele (2002) indicated that the internal
4 O3 concentration may not be zero, as previous assumed. Moldau and Bichele (2002) permitted
5 leaves ofPhaseolus vulgaris L., which have stomata on both upper and lower leaf surfaces, to
6 take up O3 at a high rate for 3 to 5 min. Exposure of the lower leaf surface resulted in up to 5%
7 of the O3 that was taken up to be diffused through the leaf, emerging from the stomata on the
8 upper surface. This suggested above-zero concentrations of O3 in the intercellular leaf air
9 spaces. The descriptive calculations and plots of Moldau and Bichele (2002) indicate that the
10 rise in internal O3 level (for both concentrations of external O3) within the first few minutes of
11 exposure is due to its reaction with an antioxidant, most probably absorbate, within the
12 apoplastic space of the leaf (Figure 9-5). The rate of rise is probably due to more complete
13 penetration of O3 with a concurrent depletion of the external antioxidant. The rise peaks at about
14 2 min for 0.88 ppm and 3 min for 0.34 ppm and then falls to a lower level. This may be due to a
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160
140
120
o
E
c
0)
o
N
O
100
80
60
03 Exposure
D 0.335
• 0.821
I
I
234
Time (minutes)
40
38
g
36 ^
"
CD
32 §
N
O
30
28
Figure 9-5. The change in the O3 concentration inside a leaf with time. Data are from O3
exposures at two different concentrations.
Source: Derived from data in Moldau and Bichele (2002).
1 replenishment of the antioxidant. The authors saw no injury to the plasmalemma (as measured
2 by penetration of a dye) and no change in the stomatal conductance for the lower concentration
3 of O3 (Moldan and Bichele, 2002). The higher level (0.88 ppm) caused the plasmalemma of the
4 mesophyll cells to pass a dye and a slight decline in stomatal conductance resulted at about
5 2.5 minutes. These data suggest that the antioxidant hypothesis is correct.
6 Gaseous pollutants flow from the substomatal cavity within the leaf into the cell, through
7 the cell wall. It is suspected that the internal concentration of the pollutant is not uniform within
8 the cavity. From within the wall, an equilibrium between the gas and aqueous phase must occur
9 at the interface where the gaseous species dissolve into the water according to Henry's Law
10 (Heath, 1980, 1987; Wellburn, 1990). It is important to understand exactly how much O3 could
11 move into the tissue of the leaves. Calculations in Table 9-6 give an indication of the amount of
12 O3 which may end up near the surface of cells within the leaf. The calculation is done for a
13 standard temperature (25 °C), an ambient concentration of O3 (0.10 ppm), and for nonspecific
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Table 9-6. The Flow of Ozone into a Leaf and Possible Reactions
The level of O3 in the atmosphere is chosen to be close to a standard and yet make calculations to other amounts
easy. The same concept will be used for all standard parameters for these calculations.
DESCRIPTION
The atmospheric level of O3 is given as:
For an air temperature of:
The perfect gas law (pV = nRT) is used to convert the O3
level into standard mks. Further, the volume for a mole of
gas (V0 = 22400 m3) will be used, from the perfect gas law
with R= 8.3144
Thus, the concentration of O3 within the atmosphere is:
O3ax 10-6-(Ta + 273.18)
°3 V0 x 273.18
The stomatal conductance of the gas must be chosen to be
standard but adjustable. The number should be as large as
typically measured, but allow for easy conversion, if
necessary. For a stomatal conductance of:
The amount of O3 that will penetrate inside the leaf (for a
typical concentration of nearly zero inside the leaf), is:
°3T
18 gsvvv
48 100
In terms of amount of water within the leaf we can assume
that about 85% of the weight is water and the density of
water is 1 g/ml. A typical leaf has a wet weight/area:
Thus, the square surface area of the leaf will translate into
water space (for concentration of chemicals), as:
ArT .=
FWL x 0.85
100
VALUES
O3a = 0.1ppm
Ta = 25 °C
= 4.873 xlo~12 moles/m3
= 1 cm/s
O3L = 2.984 xlo~14 mol/(m2 s)
FWL = 30 mg/cm2
Ar, = 0.255 L/m2
The maximum amount of toxic compound that will be
generated, assuming all the O3 is converted, is given
below. Here the units of the leaf area weight are
converted into the mks system and the water space units
are converted into L, such that the concentrations
calculated will be in mol/(L hr). The final units assume
that the O3 is present (and no back reactions occur) for one
hour (short but typical units of exposure).
O3 =(O3 /ArL)x3600
O3Lc = 4.21 x 1Q-10 mol/(L hr)
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Table 9-6 (cont'd). The Flow of Ozone into a Leaf and Possible Reactions
Thus, the maximum amount of toxic chemicals generated
per hour in a leaf would be: O3Lc = 0.42 nmol/(L hr)
Possible errors in these calculations (aside from the input numbers) are: (1) the O3 within the leaf does not react
uniformly within the leaf space, (2) the O3 within the leaf does not totally convert to any one species, (3) varied
products of O3 react leading to innocuous chemicals, and (4) O3 reactions can be catalytic and generate more by
radical reaction cycling.
1 leaves. For example, 0.3 ppm would be the same general numbers but multiplied by 3.
2 Similarly for more closed stomata, the value of 1.0 cm/s (equivalent to about 400 |imole~2-leaf
3 area s"1) for a conductance would be reduced and the smaller values would lead to a smaller
4 amount of O3 moving into the tissue. Nonetheless, these values give some indication of what
5 sort of chemical concentration can be expected. Under these conditions, a delivery rate of O3
6 into the substomatal cavity near the spongy mesophyll tissue of about 0.42 nmol/(L hr) appears
7 to be reasonable.
8
9 Process 2. Ozone diffuses into the leaf air spaces and reacts either with varied biochemical
10 compounds (path 1) which are exposed to the air or is solubilized into the water lining the cell
11 wall of the air spaces (path 2). As shown in Figure 9-6, each reaction has the possibility of
12 transforming O3 into another chemical species (a toxicant) which, in turn, may react with other
13 chemical species and lead to a cascade of reactions.
14 Within the stomata, gases react with the water at the cell's surface and generate new
15 species with the components within the cell wall region. The possible varied pathways are
16 depicted in Figure 9-7. Although these chemical reaction are poorly understood, some of the
17 fundamentals are known (Heath, 1987; 1988; Wellburn, 1990). Ozone reacts with organic
18 molecules at the double bonds to form carbonyl groups and, under certain circumstances,
19 generates peroxides, such as hydrogen peroxides (H2O2), superoxide (O2 ) and its protonated
20 form (HO2'), hydroxyl radical (HO'), and peroxy radical (HO2'). Other chemicals present in the
21 water phase can lead to many other oxygenated moities (Figure 9-6). Each of the steps are
22 generally pH dependent (Walcek et al.,1997; Jans and Hoigne, 2000).
23
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Plasma
Apoplasm Membrane
Toxicant1A .
Reaction,*
Solubilization -
4
Reaetion2A
4
Toxicant2A. -
• Reaotion1p
• Reaotiorijp
'Reaction1p-
Cytoplasm
Reaction1c
• Reaction2c
• Reaction2C-
Figure 9-6. Possible transformations of O3 within a leaf.
a.
Hydroxyl
Radical
b.
HH
Superoxide
t
^H202
Hydrogen
Peroxide
HO- H2O2
H2O
^^
Peroxyl
Radical
Figure 9-7. Possible reactions of O3 within water, (a) Ozone reacts at the double
bonds to form carbonyl groups, (b) Under certain circumstances,
peroxides are generated.
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1 Sulfhydryls are particularly easy targets, with the formation of disulfide bridges or sulfones
2 (Mudd and Kozlowski, 1975). In water, the reactions become more confusing, but some
3 products have been described by Heath and Castillo (Heath and Castillo, 1987), such as H2O2,
4 HO', and O2 (Figure 9-7). Effective detoxification reactions can occur here via antioxidant
5 metabolites and enzymes, such as ascorbate, glutathione (GSH), and superoxide dismutase
6 (SOD), if they are present at high enough concentrations (Matters and Scandalios., 1987;
7 Castillo et al., 1987; Fong and Naider, 2002). If the levels are low, it is believed that stimulation
8 of their production is a response to O3, albeit a slow one (Harris and Bailey-Serres, 1994).
9 Certainly it is possible that chemical modification of wall-specific biochemicals (Castillo et al.,
10 1987), such as glucan synthase (Ordin et al., 1969) and diamine oxidase (Peters et al., 1988), is
11 possible.
12 Process 3. Movement of reaction product(s) into the cell and enzymatic or chemical
13 transformations within the cell. It is believed that the initial site of O3 injury is near and within
14 the plasma membrane. Certainly, membrane functions, such as membrane fluidity (Pauls and
15 Thompson, 1980), permeability (Elkiey and Ormrod, 1979), K+-exchange via ATPase reactions
16 (Dominy and Heath, 1985) and Ca2+ exclusion (Castillo and Heath, 1990), are changed. The
17 similarity of wounding responses (Langebartels et al., 1991) and O3-induced membrane
18 disruption suggests the induction of normal wound-regulated genes (Mehlhorn et al., 1991;
19 Sandermann, Jr., 1998). This implies that O3 can react with components of the cell wall
20 connected to the cytoplasm through the cell wall and membrane by membrane-specific proteins
21 not directly linked to transport.
22 Ozone is soluble in water and once having entered the aqueous phase, it can be rapidly
23 altered to form oxidative products that can diffuse more readily into and through the cell and
24 react with many biochemicals. Again, the presence of internal antioxidant would be critical to
25 reduce the concentration of most oxidants. A toxic product of O3 may migrate through the
26 cytoplast to react with photo synthetic processes, or a spurious signal generated at the membrane
27 may affect some control process or signal transduction pathway (Schraudner et al., 1998;
28 Overmyer et al., 2000, 2003; Rao et al., 2000a, 2002; Sandermann, 2000; Rao and Davis, 2001;
29 Leitao et al., 2003; Vahala et al., 2003; Booker et al., 2004).
30
31
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1 9.3.3.1 Possible Reactions Within the Leaf
2 Ozone can react with many compounds within the substomatal cavity of the leaf1 to
3 produce a variety of oxidizing and toxic chemicals. Some of the possible reactions which will
4 generate H2O2, HO', and SO2 , as well as charged O3 intermediates are indicated in Figure 9-8.
5 Many of these complex reactions have been studied within water solutions through research of
6 O3 induced water purification and are very dependent upon solutes present with the solutions,
7 including FT (see von Gunten (Von Gunten, 2003)). An important point is that in alkaline
8 media, O3 forms H2O2, but in acid media O3 is relatively stable in the absence of free metal ions.
9 The rates of reaction of O3 with several important compounds, including those with a
10 double bond, the so-called Crigee Mechanism shown in Figure 9-8, can be calculated from the
11 reaction coefficient as given by Atkinson (1990) (Table 9-7). The double bond of the ascorbate
12 molecule is particularly sensitive to O3 attack. An unstable ozonide product is formed and then
13 accelerates the breakage of the double bond, leading to the formation of two products because of
14 the ring formation of the ascorbate molecule. These products are relatively unstable and can
15 lead to further reactions not shown in Figure 9-8. The rates of reactions can be calculated
16 (Heath, 1987) from the concentrations of O3 that should occur within the wall regions (Table
17 9-6). At a local concentration of 25 jiM O3, it would take 5000 s (83 min) for all of the O3 to
18 react if there was no further flow of O3. Clearly, O3 does not react rapidly with the compounds
19 in Table 9-7 and, although some of the products would be formed through the Crigee Mechanism
20 (see Figure 9-8a), they would be low in concentration2. While other radicals, such as hydroxyl
21 radical (see Figure 9-8b) can attack double bonds, the products differ. Of particular note for
22 later discussion, is the reaction of O3 with ascorbate (see Mudd, 1996; Figure 9-8b), which will
23 cleave the double bond in the ring. Unfortunately, little work has been done to characterize
24 possible products within the leaf (but see next section).
25 In a paper discussing the stability and reactivity of O3 in the pulmonary air/tissue
26 boundary, Pryor (1992) calculates that O3 has a half-life of about 7 x 10~8 s in a bilayer.
lrThe volume of the substomatal cavity (that are within the leaf immediately below the stomata) must be
regarded as the region in which most O3 reactions occur. That volume, at a relative humidity of near 100%,
possesses many diverse surfaces with varied bonding, which could alter the fate of O3.
For example, hydroxylmethyl hydroperoxide would be expected to be formed by the reaction of O3
with ethylene and its effects have been tested on peroxidases (Polle and Junkermann, 1994). Unfortunately,
the concentration of required for inhibition is much higher than would be expected to be formed within the leaf.
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a.
N0
Crigee n
Mechanism / \
;=CH2 »- 9 P
H2C — CH2
OH- + H2C=CH2
H2C=CH2
OH
\
H2C —
ON02
H2C —
H
HC=0
0
ii
HC-OH
b.
CH(OH)CH O2H
CH(OH)CH O2H
HoO
2u2
HO OH
\
O=C CH(OH)CHO2H
CHO , CHO
Further Oxidation
Figure 9-8a,b. The Crigee mechanism of O3 attack of a double bond, (a) The typical
Crigee mechanism is shown in which several reactions paths from the
initial product is shown, (b) Typical reaction of ascorbic acid with O3.
Source: Adapted from Mudd (1996).
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Table 9-7. Some Rates of Reaction of Ozone With Critical Biochemicals
[a] Double bond reactions. The second column is taken from Atkinson (1990) and transformed into Column 3.
Those rate coefficients are used to calculate the rate of reaction at a concentration of 10 ppm for the organic and
0.1 ppm for O3 in the air stream within the leaf (localized concentration of about 25 mM, see Table 9-6).
Compound x 10~18 cm3/molecules s"1 Rate coefficient (L/mole s"1) Rate of reaction (M/s)
Ethane 1.7 1.02 x 103 4.3 x I0~n
Propene 11.3 6.80 x 103 2.8 x lO"10
1-butene 11 5.91 x 103 2.5 x 10"10
trans-2 Butene 200 1.20 x 105 5.0 x 10"9
a-pinene 85 5.12 x 104 2.1 x lO"9
[b] Possible Oxidative Species. Another possibility is given by the reactions below from Walcek et al. (1997).
Reactions Rate constants
(1) O3 + OH + H2O -> H2O2 + O2 + OH kj = 3.67 x 10 mokT1 L s~'
(2) O3+ O2 -> HO + 2 O2 + OH k2= 1.26 x 109 mole"1 Ls"1
(3) O3+ HO2 -> HO + O2 + O2 k3 = 2.09 x 106 mole"1 L s"1
[c] Possible Concentrations of Other Oxidative Species. Table from Heath (1987). Based upon 100 ppm O3
in gas stream
Species
Superoxide Radical (O2')
Ozone Radical
Protonated O3 radical (HO3')
Concentration
pH7
8.75 x 10~15
4.16 x KT15
1.48 x 1Q-16
(M)
pH 9 Molecules within wall
1 x 10~12 5.5 x KT6 6.3 x 10~4
5 x 1Q-14
1 x IQ-18
Number of molecules within apoplastic space of (10~12 L) at 0.1 ppm O3.
1 However, the transit time through the lung lining fluid layer is about 2 x 1CT6 s based upon a
2 reasonable estimate for the diffusion of O3. This means that O3 would suffer nearly 29 half-
3 lives3 in passage through the layer, which would reduce O3 to about 3 x 1CT9 of the original
4 concentration — zero for all practical aspects. In the same publication, Pryor points out that any
5 sulfhydryl or ascorbate would interact strongly with O3, further reducing its net concentration.
3Here a half-life is the time that it takes the reactive species to travel a distance in which it loses 50% of its
initial concentration. Therefore for a 29 half-life, the concentration has been reduced by 2~29 or about a 10~9 decline.
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1 The reactivity of cysteine is 109, while the reactivity of tryptophan, methionine, polyunsaturated
2 fatty acids, and tyrosine is about 2 x 106, and that of phenylalanine is only 103. These numbers
3 are similar to what has been found for O3 reactivity with amino acids and proteins in aqueous
4 solutions. In glycophorin (Banerjee and Mudd, 1992) and cytochrome C (Mudd et al., 1996,
5 1997b) in aqueous solutions, only the methionine was oxidized by O3, producing sulfoxide.
6 In other proteins lacking methionine, tryptophans were oxidized only if they were in an exposed
7 position on the surface of the proteins (Mudd et al., 1997b). Treatment of red blood cell ghosts
8 with O3, oxidized peripheral proteins of the plasma membrane before it oxidized lipids (Mudd
9 etal., 1997a).
10
11 9.3.3.2 Toxicants Within the Wall Space
12 While Mehlhorn et al. (1990) is often thought to have shown that free radicals were formed
13 in plant leaves under O3 exposure, careful reading of that paper clearly shows that there was no
14 real evidence of free radicals induced by O3. Living tissues have many free radical signals,
15 making it difficult to observe changes in free radicals. Further, the work of Grimes et al. (1983)
16 has also been cited as showing the presence of free radicals in living tissues due to O3 exposure;
17 however, no radical signals were found unless certain organic acids (e.g., caffeic acid) were
18 added to the tissue with the O3 exposure. They used the radical trap TMPO (tetramethylphrrolise
19 1-oxide) which reacts with many types of free radicals to form a stable radical that can be used
20 to "trap" or increase the amount of radical present (see Figure 9-9a). Ozone would directly react
21 with this trap only if it were bubbled into the solution, not passed over the top of the solution.
22 In the presence of sorbitol or caffeic acid, the trap would indicate the presence of OH radical,
23 which would mean that O3 ->• HO'. Superoxide dismutase, catalase, or EDU had no effect upon
24 this signal, suggesting O2 and H2O2 were not involved in the above sequence. Both O3 and O3
25 plus caffeic acid had no effect upon the protoplasts' intactness or viability. Thus, 10~5 M HO'
26 and/or 0.30 to 0.40 ppm O3 did not react with the cell membrane. They found no signal in
27 normal cells after subjecting the leaf to O3 and concluded that the radicals were produced via a
28 concerted mechanism with the acid. This does not fit with the mechanism postulated by
29 Mehlhorn et al. (1991), which involved a reaction of wound-induced ethylene and O3 at the wall
30 level to generate some free radicals.
31
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a.
ESR spectra from solution treated
with O3 at 0.3 pg/g dissolved for
60 s (1a) control, (1b) with caffeic
acid; spin trapped: DMPO
(Grimes et al., 1983)
ESR spectra from bean
(Phaseolus vulgar!s cv. Pinto)
leaves pretreated with spin
trapped: PEN after a 4-h
fumigation with O3 (0-300
nl/L)(Mehlhornetal., 1990).
C.
EPR signal in white light before
(broken line) and after (dotted
line) exposure of a bluegrass
leaf to 1000 ug rrf3 O3 in the
EPR cavity for 1 h. The solid line
depicts the difference signal.
(Runeckles and Vaartnou, 1997)
Figure 9-9. Varied ESR radicals, trapped and not, generated by ozone under somewhat
physiological conditions, (a) The generation of a DMPO-trapped radical with
caffeic acid in water solution (Grimes et al., 1983). (b) The generation of a
DMPO-trapped radical within bean (Phaseolus vulgaris cv. Pinto) exposed to
0.10 ppm O3 for 4 hours. The lower trace is the ESR signal produced with
0.3 ppm O3 (Mehlhorn et al., 1990). (c) The EPR signal produced within a
bluegrass leaf exposed to ppm of O3 for 1 h (Runeckles, 1997). Although no
trapping agent was used in this experiment, the signal is complex because of
various free radicals normally present within the illuminated leaf.
1
2
3
4
5
The hypothesis that the production of wound-induced ethylene by O3 exposure and its
reaction with O3 would result in the production of radicals was tested by Mehlhorn et al. (1990),
using electron paramagnetic resonance spectroscopy. After 4 h of 300 ppb, an EPR signal of a
compound was detected which resembled a butonyl radical (Figure 9-9b). Using 70 ppb, the
signal was reduced by about one third of an ethyl radical4, leading to injury. However, the
spraying of the plant with 1-aminoethoxyvinyl-Gly (AVG), which reduces the production of
4The reaction would be: O3 + H2C = CH2 -»• varied C-l compounds, due to double bond cleavage, at a
rate constant of 1.7 x l(T18 cnf/molecule sec = 1.02 x 103 NT1 s"1 (Atkinson R., 1990). This should be compared
with a reaction of the hydroxyl radical with ethylene, which has a rate constant of 8.52 x 10~12 cm3 molecule"1 s~',
or 106 x faster.
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1 ethylene and visible injury, had no effect upon the EPR signal, suggesting that the radical is not
2 on a direct sequela to visible injury.
3 Runeckles and Vaartnou (1997; Figure 9-9c) discovered a signal by subtracting other EPR
4 signals of the leaf, which seemed to be due to an O3 reaction with plant material, using 0.48 ppm
5 O3. This difference signal looked very much like O2 . At a lower concentration, they observed
6 that this signal still occurred but accumulated more slowly. Both bluegrass and ryegrass leaves
7 seemed to saturate after about 5 h of exposure at 22 to 28 units of signal, while radish leaves
8 reached a maximum of 7 units at 3 h and then declined. The problem, which is typical of any of
9 these methods, was that the detached leaf had to be rolled and placed into the EPR detection
10 cavity. Reichenauer et al. (1998) also detected an undefined free radical signal that seemed to be
11 related to a Mn(II) spectrum. The Nandu and Perlo cultivars of wheat were more sensitive to O3
12 than Extradur (according to growth rate and closure of stomata under an O3 exposure of 80 ppb
13 for 8 h/day, 7 days/week over 100 days), and these more sensitive cultivars had a greater, but
14 insignificant (P = 15%) EPR signal. Thus, data showing any production of a free radical must be
15 approached with some skepticism.
16 With an O3 delivery rate of about 25 |iM/h (Table 9-6), only 250 jiM would be found after
17 a full day, if all of the O3 were stable. While the use of free radical traps is the best method
18 available to observe any build-up of radicals, the traps are not as specific to individual radicals.
19 Currently, studies should be looking for hydroxyl radicals, superoxide, hyroperoxides, ethylene
20 radicals, and ascorbate radicals.
21
22 9.3.3.3 Products of Ozone
23 Ozone should reach a certain concentration in the substomatal cavity, which is dependent
24 upon its entry speed and its reactivity with the wall constituents. Once near the apoplastic space,
25 O3 moves in two different pathways (Figure 9-6). It can react with constituents that are within
26 the wall as a gas in reaction 1A (path 1); or it can solubilize into a water space and travel to
27 another region within the water space and react through reaction 2A (path 2).
28
29 Hydrogen Peroxide
30 Hydrogen peroxide, until recently was thought to be purely a toxic compound for cells.
31 However, it is now clear that it functions as a signaling molecule in plants and mediates
January 2005 9-48 DRAFT-DO NOT QUOTE OR CITE
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1 responses to abiotic and biotic stressors (Figure 9-10). Generation of H2O2 is increased in
2 response to various stressors, implicating it as key factor mediating the phenomena of
3 acclimation and cross tolerance, in which exposure to one stressor can induce tolerance of
4 subsequent exposure to the same or different stresses (Neill et al., 2002). The signaling response
5 to attack by invading pathogens using H2O2 has been described (Mehdy, 1994; Simon-Plas et al.,
6 1997). The reactions leading to hypersensitive cell death are caused by a pathogen recognition
7 step (Figure 9-10a), probably due to the plant cell wall releasing oligosaccharides in response to
8 the pathogen enzymatically breaking down the cell wall to penetrate it. A feed-forward step in
9 which H2O2 increases the level of benzoic acid leads to the activation of the hydroxylase step in
10 the production of salicylic acid and to a feed-back step in which the salicylic acid increases the
11 production of H2O2 (Leon et al., 1995).
12 An elicitor, e.g., a bacterial or fungal pathogen, induces a cascade of reactions within a cell
13 (Figure 9-1 Ob). Some of the lipid reactions are thought to be due to the opening of the Ca2+
14 channels and the alkalination of the cell wall region. The oxidative burst due to H2O2 production
15 is believed to lead to the transformation of a small population of lipids into jasmonic acid, which
16 is a secondary messenger.
17 Hydrogen peroxide also has an oxidative role in lignification (Schopfer, 1994). In the
18 interaction of lignification and the beginning processes of hypersensitivity, pectinase produced
19 by the pathogen disrupts pectin and dissolves the cell wall. Fragments of the dissolved cell wall
20 trigger an increase in the transcription of peroxidases within the remaining cell wall, leading to
21 lignification, which is a cross-linking of the cell wall that does not use pectin. This prevents
22 further pathogen disruption of the wall and reduces its further entry into the plant cell.
23 It is believed that the first species generated through a one-electron reduction of molecular
24 oxygen is SO2 . That generation is carried out using a cytochrome b6 by the NAD(P)H oxidase
25 located on the cell membrane (Auh and Murphy, 1995). In the acid region of the cell wall, SO2
26 is converted by a protonation and dismutation to H2O2. The induced oxidative burst is believed
27 to play a role in stimulating the Cl" and K+ efflux, generating an alkalinization of the
28 extracellular space (Cazale et al., 1998). In the wall region, H2O2is not especially toxic, as no
29 necrosis was reported in tobacco when 500 mM peroxide was infiltrated into the leaf tissue.
30 However, the production of salicylic acid and benzoic 2-hydroxylase can be induced with only
31 30 and 0.3 mM H2O2 respectively, indicating some metabolic signaling (Leon et al., 1995).
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a.
Pathogen Recognition
I
Free Radical Generation
H2O2 Generation
I
Accumulation of Benzole Acid
^ Benzole Acid
2-Hydroxylase Activation
I
— Salicylic Acid Accumulation
I
Systemic Salicylic Acid Increase
Defense Activation
Hypersensitive Cell Death
Local Resistance
Global Resistance
/ell Membrane C
Ascorbate
Peroxides
OH" \
1 *
\ / „ .(
G Proteins Open Ca2+ Phosphorylation * (
(Rac2/Rap1A) Channels Cascade %
I
Pathogen
/ Kill
/ t i
•( "Oxidative
+ Burst
Cell Wall
->. Structura
Changes
/ NADPH Hydroperoxides
H+ Oxidase ./
Jasmonic
/ Acid
Influx
Signal Transduction
Transcription
Defense Genes
Figure 9-10. Pathogen-Induced Hypersensivity. (a) The reactions leading to
hypersensitive cell death and the formation of a global response of salicylic
acid, (b) The cascade of the elicitor-induced reactions within the cell.
1
2
On the other hand, 1M H2O2 infiltrated into soybean will generate lipid peroxidation after 1 h
with a peroxidation rate of 15 nmol/g-FW h (Degousee et al., 1994). Cells react to the system5
Soybean suspension cells were inoculated with Pseudomonas syringae pv syringae, which generate an
active oxygen response. Light emission by luminol, reacting with H2O2, was the assay for the peroxide.
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1 and generate peroxide scavenging compounds within 1.5 to 2 hours, which appear to "mop up"
2 the excess H2O2 (Baker et al., 1995).
3 After O3 exposure in birch, H2O2 has been found in the wall (Pellinen et al., 1999).
4 By using CeCl2 as a cellular stain for H2O2 (as a cerium perhydroxide precipitate), Liu et al.
5 (1995) observed a gradual development of stain after 8 h of O3 exposure (at 150 ppb). After 2 h
6 exposure, H2O2 stain was visible on the surfaces of both sets of mesophyll cells. Accumulation
7 of H2O2 stain continued for 16 h after exposure, suggesting a triggered-reaction rather than O3
8 decomposition itself. H2O2 stain was present in the mitochondria, peroxisomes, and cytoplasm,
9 but not in the chloroplast. If methyl viologen (MV) was given to the leaves and then the leaves
10 were exposed to light, H2O2 stain could be observed within the chloroplast. This indicated that
11 the stain worked within the chloroplast if H2O2 were generated by the Mehler reaction
12 (MV+O2 ). Thus, apparently, for birch, O3 exposure does not generate excess H2O2 within the
13 chloroplast. Furthermore, these sets of experiments indicate that O3per se does not generate the
14 H2O2, but rather triggers stress-related H2O2 formation similar to what occurs in a pathogen
15 attack (the Reactive Oxidative Species or ROS reaction).
16 The presence of higher than normal levels of H2O2 within the apoplastic space is a potential
17 trigger for the normal, well-studied pathogen defense pathway. Figure 9-1 Ob depicts such a
18 pathway and suggests that all the events and activation of pathways/genes caused by pathogen
19 defense could be observed when plants are fumigated with O3. These events in Figure 9-1 Ob will
20 be alluded to in later sections.
21 H2O2 has been linked to the hormone ABA-induced closure of the stomata by activating the
22 calcium influx in guard cells (Pell et al., 2000). The addition of H2O2 at a level of only 5 mM to
23 a guard cell preparation will cause a dramatic increase (ca. 9*) in electrical current at the
24 hyperpolarizing potential of-200 mV. Amounts as low 50 mM H2O2 will cause a less but still
25 sizable increase. Membrane stability is unaffected by the H2O2 and the activation of the channel
26 requires only about 2 to 3 minutes. Pell et al. (2000) also found that ABA induced the
27 production of H2O2through ROS accumulation (also see Zhang et al., 2001).
28 Certain levels of ABA within the leaf lead to stomatal closure. The inactivation of a
29 phospho-tyrosine-specific protein phosphatase (ABI2) is an inhibitor of stomatal opening
30 induced by ABA but that enzyme is inhibited by H2O2 (Meinhard et al., 2002). This means that
31 H2O2 shifts the sensitivity of the stomatal opening to ABA (Figure 9-11), making the stomatal
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ABA
Level
9s
Control
Stomatal
Closure
[ABA]
Figure 9-11. The interaction of H2O2 and Ca2+ movements with ABA-induced stomatal
closure. It is well known that certain levels of ABA within the leaf lead to
closure of the stomata within the leaf. That level, however, can be shifted to
make closure more or less sensitive to a given level of ABA. Recently it has
been shown that H2O2 (externally or produced by the plant) within the cell
wall region can shift that sensitivity. Here ABA stimulates the production of
H2O2, which in turn increases the rate of Ca2+ moving from the wall region
into the cytoplasm. That shift in internal Ca2+ level increases the closure of
the stomata. Hydrogen peroxide also blocks the activation of a polypeptide
(ABI2) that inhibits stomatal closure seemingly induced by ABA.
Source: From Assmann (2003); Assmannand Wang (2001); and Zhang et al. (2001).
1 complex more sensitive to ABA. Thus, for a given level of ABA present in the guard cell
2 complex due to environmental factors (e.g., low humidity, high air temperature, or low soil water
3 potential), the generation of H2O2 would (by inhibiting ABI2) induce a closure of the stomata by
4 increasing the sensitivity of the guard cells to ABA. In the past, it has been difficult to
5 understand why O3 would often decrease conductance in some cases, but not always (Heath,
6 1994b). This interaction between H2O2 and ABA could help understand this complexity.
7
8 Ethylene Reactivity
9 Ethylene (ET) is produced when plants are subjected to biotic stressors e.g., attacks by
10 insects, fungi, and bacteria or abiotic stressors such as wounding or environmental stressors such
11 as heat, cold, or oxidative stress and O3. If an O3 stress has induced a wounding response with
12 ET release, then ET within the substomatal cavity could react with O3, generating some
13 relatively noxious chemicals (see Figure 9-6). The relationship between O3 injury and wounding
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1 is supported by the observation by Mehlhorn et al. (1991) that an inhibitor of ET formation,
2 AVG (an inhibitor of ACC synthase, a committed step to ET production), would block ET
3 formation and inhibit visible injury. Other studies with polyamine (which is closely linked to
4 ET production), including those of Ormrod and Beckerson (1986) who fed polyamines to the
5 transpirational stream and prevented visible injury, suggested a close involvement of both
6 pathways to the production of visible injury. Both the lack of ET production or an increased
7 level of polyamines slowed or prevented visible injury.
8 This concept was taken another step by Langerbartles (Langebartels et al., 1991). The
9 linkage to the pathogen wound responses and visible injury is well established (Sandermann,
10 1996). Sandermann (1998) used a system of Bell B and W3 tobacco, plants with differential O3
11 sensitivities, in which the O3 exposure level was chosen such that the sensitive cultivar was
12 injured, while the tolerant one was not. This led to a marvelous control which could be used to
13 their advantage. They followed a time sequence to show that the rise of varied systems followed
14 the same order as for a pathogen attack (Heath, 1994a).
15 More recent studies, however, indicate that O3 responses resemble components of the
16 hypersensitive response (HR) observed in incompatible plant-pathogen interactions (Sanderman
17 et al., 1998). The similarity to the HR response may be related to the occurrence of ROS in the
18 apoplast. The O3-derived ROS apparently trigger an oxidative burst in the affected cells by an as
19 yet unknown mechanism. An oxidative burst is similar to one of the earliest responses of plants
20 to microbial pathogens and is an integral component in HR-related cell death (Overmyer et al.,
21 2000).
22 In plants exposed to O3, ET synthesis is a result of the specific ET induction of the genes
23 encoding 1-aminocyclopropane-l-carboxylase synthase (ACS), one of the fastest and most
24 obvious responses to O3, which has been mechanistically linked to the regulation of O3 lesion
25 formation. Biosynthesis of ET inhibited with ACS inhibitors significantly reduced the induction
26 of lesion formation in plant leaves exposed to O3 concentration (Vahala et al., 2003; Melhorn
27 and Welburn, 1987; Mehlhorn et al., 1991). Ethylene biosynthesis correlates best with O3
28 (Vahala et al., 2003; Overmyer, 2000). These data support the concept that elimination of ET
29 formation will prevent visible injury.
30
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1 Ethylene-Interaction with Injury and Conductance
2 Increased ET production by plants exposed to O3 stress was identified as a consistent
3 marker for O3 exposure decades ago (Tingey et al., 1976). They exposed more than 20 plant
4 species and cultivars to O3 to determine whether the production of O3-induced stress-ET could be
5 used to determine differences in plant sensitivity to O3. Their studies suggested that increased
6 production of stress-ET correlated well with the degree of foliar injury that developed within
7 hours or days after O3 exposure. The amount of ET released was exponentially related to the O3
8 exposure. Furthermore, the amount of O3-induced ET declined with repeated exposure,
9 indicating an acclimatization to O3. This acclimatization effect associated with repeated
10 wounding has not yet been well described. The release of wound-induced ET is not linear with
11 time, but declines after the initial response (Stan et al., 1981), as is also seen after O3 exposure
12 (Stan and Schicker, 1982). The stress-induced ET production correlates better with O3 exposure
13 level than with exposure duration. In other words, peaks of high O3 (rather than accumulated
14 dose) generate a higher rate of ET release, at least for a single O3 exposure under an acute dose.
15 The production of ET after an O3 exposure is thought to be a typical wounding response
16 (Tingey et al., 1975). Prevention of ET release may prevent the formation of visible injury
17 (Mehlhorn and Wellburn, 1987). However, the question arises as to whether this effect was
18 limited to the prevention of visible injury or if the chemicals used to prevent ET release closed
19 the stomata. Using Glycine max L., Taylor et al. (1988b) showed clearly that AVG did not
20 necessarily close stomata nor inhibit carbon assimilation/>er se.
21 The correlation of ET release with O3-induced visible injury was likewise shown in pea
22 cultivars (Dijak and Ormrod, 1982). With O3 exposure (generally 6 h at 0.3 ppm), the stomata
23 closed by -50% within 3 hours after a dose of 3 x 10~5 mol cm"2 (with an average rate of
24 2 x io~9 moles cm"2 s"1, as calculated from their data). Both sensitive and insensitive cultivars
25 had a visible-linked-injury ET release, but sensitive cultivars scored higher both in visible injury
26 and in ET release after a given exposure.
27 Gunderson and Taylor (1988, 1991) used exogenous ET to alter the gas exchange of
28 Glycine max and found an exponential, but not simultaneous, decline of both stomatal
29 conductance and carbon assimilation with ET. Interestingly, the exogenous ET caused a slight
30 rise in difference of CO2 within and without the leaf, indicating a lowering of internal CO2,
31 which was not observed in the experiments of Farage et al. (1991) for O3 exposure. Ethylene
January 2005 9-54 DRAFT-DO NOT QUOTE OR CITE
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1 does inhibit both stomatal conductance and carbon assimilation to some extent (Taylor et al.,
2 1988b). Thus, one could postulate that O3 generates a wounding response with a production of
3 ET, which would, in turn, generate the change in stomatal conductance and photosynthesis.
4 Clearly, these multiple events may have confounded some earlier studies.
5
6 9.3.3.4 Antioxidants Within the Apoplastic Space
7 The first line of defense against O3 is a closure of the stomata to exclude its uptake. This is
8 counter-productive for efficient photosynthesis, but some amount of closure limits the rate of O3
9 deposition into the leaf tissue to allow for a secondary line of defense to detoxify the O3. The
10 secondary line of defense involves a range of antioxidants, which are highly reactive to the types
11 of chemicals that can be generated by O3. Several antioxidant proteins are stimulated by O3
12 in Plumbagini folia, including glutathione peroxidase (GSH-PX), SOD, and catalase. The
13 timescales for changes in their levels vary: some rise rapidly, while others rise more slowly.
14 The pattern of changes in these particular proteins varies greatly among different species and
15 conditions.
16
17 Ascorbate Within the Cell Wall
18 Most of the recent reports indicate that ascorbate within the cell wall is the real first line of
19 all defense. Ascorbate within the wall declines when the tissue is exposed to O3 (Luwe et al.,
20 1993; Moldau, 1998; Turcsanyi et al., 2000; Zheng et al., 2000). This decline appears to be
21 closely linked to the amount of O3 penetrating the leaf tissue.
22 It has long been suspected that intracellular antioxidants play a role in preventing O3-
23 induced injury to plant cells. Variation in the types of biochemical compounds present in the
24 apoplastic space can give rise to a multiplicity of reactions with O3, but the predominant
25 biochemical species is ascorbate. Ascorbate is water-soluble, present in the solution where O3
26 can dissolve, and is highly reactive. Unfortunately, a variety of antioxidants are found
27 throughout the cell and any measurements of one particular type within the total leaf tissue can
28 give misleading results. For example, ascorbate is present within the cell wall, within the
29 cytoplasm (Moldau, 1998; Burkey, 1999), and within the chloroplasts (Law et al., 1999); and
30 ascorbate can move between the cytoplasm and the cell wall with relative ease (Figure 9-12;
31 Bichelle et al, 2000). The total of all ascorbate pools are measured when the tissue is ground and
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Cell
Apoplasm Membrane Cytoplasm
a.
Ma late
NAD
NADH
> De hydro-.
ascorbate
Ascorbate
HI Cytb
Dehydro-
ascorbate
-^/AscoTbateJ
II trai sporter(s) \(
De hydro- A
ascorbate
Dehydro-
ascorbate
[ Ascorbate
Dehydro-
ascorbate
b.
Ascorbate
Monodehydro-
ascorbate
Ascorbate
Carrier
Dehydro-
ascorbate
Ascorbate
Cytochrome b
Monodehydro-
ascorbate
NADH
Oxidoreductase
NAD+
Ascorbate
De hydro-
ascorbate
GSH
•GSSG
NADPH
NADP+
Dehydro- Glutathione
ascorbate Reductase
Reductase
Figure 9-12. The reaction of ascorbate within the apoplasm of the cell wall and its
ultimate reduction/oxidations.
(a) Movements of reducing power (from Dietz, 1997).
(b) The use of glutathione to maintain the level of ascorbate
within the cell wall region (from Horemans et al., 2000).
January 2005
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1 assayed. If the cell wall ascorbate drops by 50% due to O3 exposure but all other tissue
2 concentrations remain the same, the measurement of the total loss is dependent upon the amount
3 of ascorbate within the cell wall. Turcsanyi et al. (2000) showed that, compared to the
4 concentration of apoplastic ascorbate, the rest of the cells contained about 38 times as much.
5 So a 50% loss of apoplastic ascorbate would be converted into only 2 to 3% loss of the total
6 ascorbate.
7 The ascorbate deficient Arabidopsis thaliana mutant has proven to be a powerful tool in
8 furthering the understanding of ascorbate biosynthesis in plants (Smirnoff et al., 2001). Three
9 classes of mutants were formed when A thaliana seed was mutagenized with ethyl
10 methanesulfonate: (a) those deficient in SOD, (b) those that failed to accumulate more
11 antioxidant proteins upon increased O3 exposure, and (c) those that were deficient (but not
12 depleted) in ascorbate. The low-ascorbate mutant type had 50 to 60% less ascorbate than the
13 wild type and displayed more foliar injury. This mutant is involved with the coding of the
14 GDP-D-Mannose pyrophosphorylase enzyme6 in the Smirnoff-Wheeler pathway for ascorbate
15 biosynthesis. Smirnoff et al. (2001) also suggested that other pathways can produce ascorbate
16 without relaying upon the pyrophosphorylase step, but most probably at a slower through-put
17 rate, because any fully ascorbate-deficient mutant would be lethal, perhaps because of ascorbate
18 use as a cofactor rather than its antioxidant properties.
19 The ascorbate peroxidase (APX, which uses ascorbate to detoxify peroxides) family
20 consists of at least five different isoforms, with isozymes in the apoplastic and cytosolic space.
21 Furthermore, most forms of ascorbate can move through the plasma membrane (Figure 9-12;
22 Bichele et al., 2000), making the levels of all forms of ascorbate interdependent and able to at
23 least partially influence each other. Dehydroascorbate (DHA) can be broken down into other
24 smaller fragments easily in vivo and represents a continuous loss of ascorbate from varied parts
25 of the cell if ascorbate is allowed to remain in the oxidized form in some regions. In fact, the
26 turnover rate in leaves is estimated to be from 2 to 13% / hour, depending upon species and
27 developmental age (Smirnoff et al., 2001). There are apparently three pathways for ascorbate
6 EC 2.7.713, Mannose-1 phosphate guanylyltransferase; mannose + GTP —> GDP-mannose + ppi;
this product leads into cell wall polysaccharide synthesis and protein glycosylation through GDP-galactose and
GDP-fucose and ultimately, through Galactose, into ascorbate synthesis.
January 2005 9-57 DRAFT-DO NOT QUOTE OR CITE
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1 turnover (Figure 9-12a). The typical reaction is a reduction of DHA into ascorbate from which
2 an oxidative step generates DHA. Pathway I requires a reductive step using NADH external to
3 the plasma membrane generated from internal malate using a malate/oxaloacetate transporter.
4 Pathway II uses a direct transporter of ascrobate/DHA. Pathway III moves the electron(s)
5 required through a cytochrome b system, maintaining two separate pools of ascorbate/DHA
6 within the cytoplasm and within the wall. Each of the pathways (Dietz, 1997) represented by
7 Roman Numerals in the Figure 9-12, require only one NAD(P)H molecule to reduce the DHA
8 molecule back to ascorbate. However, the transport properties and redox potential of the cell
9 differ for each pathway. The efficiency of the reduction of DHA is dependent upon the redox
10 coupling and the region in which the chemical species is located.
11 Turcsanyi et al. (2000) exposed broad bean (Viciafabd) grown under two regimes in
12 duplicate controlled chambers: charcoal/Purafil filtered air (CFA) or (CFA) plus 0.075 ppm O3
13 for 7 h/day for 28 days (chronic exposure) or exposed to 0.150 ppm for 8 h (acute exposure).
14 Responses of the two set of plants were similar except for stomatal conductance, which was 50%
15 lower in the chronic exposure plants. Plants grown under acute exposures developed visible
16 injury, while plants grown under chronic conditions developed no visible injury. Within an hour
17 of the start of the acute exposure, the stomatal conductance was reduced by nearly 40% and
18 assimilation was reduced by nearly 18% in the clean air plants; a reduction in conductance was
19 only 21% and assimilation 16% in the plants subjected to chronic O3 exposures. The
20 assimilation was affected similarly in both cases, while the conductance showed less of a
21 percentage drop in the chronic O3-exposed plants, beginning at a lower O3 level. The similarity
22 of the assimilation indicated that the stomata were not limiting assimilation in either case before
23 acute exposure. More to the point, the decline in ascorbate in the apoplastic space due to the O3
24 exposure was ".. .more often than not, on the borderlines of statistical significance." However, a
25 30% decline in ascorbate after 4 hours of acute O3 in both cases was observed. This lack of
26 significance may be due to a relatively large standard error of the data, which in turn may be due
27 to the difficulty of extracting and measuring ascorbate from the apoplastic space in quantitative
28 terms.
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1 The chemical reaction7 of ascorbate and O3 is given by the molecular rate constant of
2 4.8 x 1071VT1 s"1. This is some 50,000x the rate constant for ET. Of course, it depends upon the
3 relative concentration of ascorbate and ET, but it is likely that ascorbate is in higher
4 concentration than ET. One would then expect that the rate reaction of ascorbate with O3 would
5 greatly dominate any possible reaction of O3 with ET. For a concentration of ascorbate in the
6 range of 1 mM and for an O3 concentration of about 0.1 ppm or 4.2 x 1CT9 M, the detoxification
7 rate would be 4.8 x 107 x i(r3 x 4.2 x l(T9 M s"1 = 2.0 x l(T4 M s"1. Turcsanyi et al. ( 2000)
8 calculated an O3 flux of about 1.6 x io~9 moles m"2 sec"1. With a wall thickness of 0.12 x io~6 m
9 and all the O3 flux going into the wall region, this would give about 1.3 x 10~2 mol m"3 s"1 or
10 1.3 x 1Q~5 M s"1 flux, which is less than 10% the detoxification rate.
11
12 Glutathione
13 Many of the initial studies of O3 exposure used high concentrations and measured only the
14 total sulfhydryl contents of the tissues. For example, in some of the earlier work, exposures of
15 tobacco to 1 ppm O3 for 30 min induced a 15% loss of the total sulfhydryls (0.74 |imole/g-FW;
16 Tomlinson and Rich, 1968). These results are similar to other studies at high O3 levels (Dugger
17 and Ting, 1970). It is now suspected that severe injury in their studies resulted in a massive
18 collapse of the cells releasing most of their internal constituents. Much of the oxidation thus
19 observed may have been the result of chemical oxidations of the O3 that subsequently entered the
20 damaged tissue. Even under milder conditions, changes in sulfhydryl components have still
21 been noted and any sulfhydryl on the surface of the cell would be at risk due to its high reactivity
22 with O3 (Mudd et al., 1969; 1997a). For example, the level of sulfhydryl compounds within the
23 protein of isolated chloroplasts declined about 66% when the chloroplasts were subjected to O3
24 (about 1 jimole O3) exposure (Mudd et al., 1971).
25 At this stage, it is important to note that there are inherent problems with metabolic studies
26 of full tissues. The first is that most organs have several different types of tissues. For example,
27 leaves have, at the minimum, epidermal and vascular tissues and two types of mesophyll cells.
28 Each type of cell may be metabolizing quite differently, producing very different levels of
7These chemical rate constants are those constants within a bulk solution. In the apoplasm, the possibility
exists for the chemicals to be preferentially oriented near a surface; so the constants may not be the same as for
bulk solutions.
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1 metabolites and enzymes. Furthermore, most pathways are well regulated and after any small
2 disruption, the pathway tends to return to near its former stability. Changes in the level of
3 enzymes are likewise difficult to measure. Many enzymes function below their maximum
4 activities. Their speeds of reactions are often increased through regulation, rather than through
5 the production of more enzyme.
6 Glutathione is a three-amino acid peptide, which has antioxidant properties due to its free
7 reducing sulfhydryl group (G-SH). Glutathione is generally kept in its reduced form by
8 glutathione reductase (GR) with the reaction:
9
GS - SG(oxidized) + 2e~ + 2H+ -> 2GSH (9-2)
10
11 GR has six isoforms8 within the chloroplast and six isoforms outside. The optimum activity
12 occurs at pH 7.8, suggesting it is located within the stroma of the chloroplast or the cytoplasm
13 rather than in the cell wall, which is at pH 4-5 (Madamanchi et al. (1992). Clearly, an increased
14 expression of GR (generated through transgenic implants) is important within the chloroplast to
15 prevent of some oxidations9 (Aono et al., 1995; Foyer et al., 1995).
16
17 Catalase
18 Catalase, even though it breaks down H2O2, does not appear to protect plants from O3
19 exposures. Two principle reasons may cause this lack of reactivity: (1) catalase has a high Km
20 for H2O2 and a low rate coefficient and (2) seems not to occur within the cell wall regions but
21 rather in the cytoplasm and peroxisomes (Buchanan et al., 2000). Only a few reports suggest
22 that catalase is increased by exposure to O3 (Azevedo et al., 1998). Unfortunately H2O2 induced
23 by some forms of wounding in mesophyll cells can lead to induction of an increase in GSH and
24 the transient production of catalase (Vanacker et al., 2000). In general, it seems that catalase is
An isoform is the same enzyme, with the same structure and perhaps within the same organelle, but its
promoter region has different DNA codes. Thus, each protein segment is induced by different signals, and so its
enzyme can be formed in response to different environments. This is in contrast to isozymes, which classically are
similarly reacting, but structurally different, enzymes in different compartments.
9Typically this protection is observed in the paraquat sensitivity of plants. In this assay, added paraquat,
the herbicide which intercepts electrons from the reducing end of photosystem I in the chloroplast, caused
oxidations, chlorophyll loss, and death due to the buildup of superoxide and peroxides.
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1 not really involved primarily in the defense of the cell due to O3 attack but rather may be a
2 secondary response. The reaction of catalase (Scandalios, 1993) is as follows:
3
4 2H2O2 -> 2H2O + O2 K = 1.7 x ICT7 N/T1 sec'1 (9 3)
6
7 Superoxide Dismutase
8 The varied compounds that O3 can produce upon entering an aqueous solution are very
9 similar to those involved in the HR when plants are infected by an avirulent pathogen (Figure
10 9-10). The sequence of the plant response to the pathogen is: (1) recognition of the gene
11 products of the pathogen by the plant (elicitor), (2) generation of an immediate phytoresponse to
12 attempt to localize the attack and its products, and (3) generation of a systemic acquired
13 resistance (SAR) to subsequent attack by the pathogen. Inducible defense responses are
14 phytoalexin synthesis and production of pathogenesis-related proteins (PR). One aspect of this
15 total response is the production of O2 and H2O2 by the cell (Lamb and Dixon, 1997). The
16 elicitor can generate a transient alkalinization of the apoplast, up to pH 7.2, caused by a lowering
17 of the H+-pump rate and a increase in the H+-influx/K+-efflux exchange. Other effects include:
18 a weak accumulation of transcripts for PAL (phenylalanine lyase); a larger and rapid induction
19 of glutathione S-transferase, GSH-PX; oxidative cross linking of cell wall proteins which is
20 blocked by ascorbate acid; generation of localized apoptosis; and rapid influx of Ca2+, which
21 activates apoptosis among other pathways (Lamb and Dixon, 1997). These effects seem to be
22 very similar to those induced by O3 exposure (Sandermann, 1996; 1998).
23 The putative antioxidant enzyme SOD (Equation 9-4 and Table 9-8) catalyzes the
24 oxidoreductase reaction, which eliminates SO2 by dismutation (Bowler et al., 1992):
25
26 2O2- + 2H+ —SOD > H2O2 + O2 K = 2.4 x 109 M'1 seer1. (9-4)
27
28
29 The number, as well as the activity, of isozymes of each type of SOD in Table 9-3 can vary
30 with plant species. However, the isozymes that have been tabulated are Cu-Zn SODs, in cytosol
31 and chloroplast; Fe-SOD, active in chloroplast stroma, and Mn-SOD, active in mitochondrial
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Table 9-8. Superoxide Dismutase Isozymes and Isoforms
Reaction: 2H+H
Isozymes
Cu-Zn
Fe
Mn
2 O2 - H2O2 + O2
M.W.
20kDa
23kDa
23kDa
Isoforms
csdl
csd2
csd3
fsdl
sd2
fsd3
msdl
Cytolocation
Plastid
Peroxisomal
Mitochondrial
Plastid
Mitochondrial
1 matrix (Karpinski et al., 1993). In the experiment demonstrating the activation of varied SODs,
2 there were three Cu-Zn SOD (csdl, csd2, csd3\ three Fe-SODs (fsdl,fsd2,fsd3), and one
3 Mn-SOD (msdl) (Kliebenstein et al., 1998). Ozone sensitivity was determined by exposure of
4 plants to 8 h of 0.33 ppm of O3. csdl induced by O3 and UV-B was one of the earliest SOD
5 increases and most pronounced responses for mRNA and protein. Also, some increase in csd3
6 (thought to be peroxisomal) was induced when the plants were exposed to a high-intensity light
7 pulse; msdl was unresponsive to the environmental stressors used here, including O3; and csd2
8 (thought to be chloroplast) showed little increase. The fsdl isozyme (present in the apoplasm)
9 showed a slight decrease. On the other hand, an early report on snap beans in which the
10 experimenters used EDU, N-[2-(2-oxo-l-imidazolidinyl)ethyl]-N-phenylurea (Carnahan et al.,
11 1978; Kostka-Rick and Manning, 1993) to prevent visible injury by O3, 4 h O3 exposure at
12 0.45 ppm was correlated with an increase in general enzyme activity of SOD, i.e., the level rose
13 nearly 2.5x in 2 weeks at a level of 50 mg EDU per pot (Lee and Bennett, 1982). It is believed
14 that EDU may induce SOD, which then protects the plant. While gross assays of enzyme
15 activity have not proven to be very useful in understanding the mechanism of O3 action, in a
16 well-crafted, long-term study involving ponderosa pine clones. Benes et al. (1995) stated that
17 "changes in antioxidant enzyme activity were not a consistent response to the O3 fumigation, but
18 when observed, they occurred most often in the O3-sensitive clone and in symptomatic,
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1 fumigated branches.. .total (intra- and extracellular) activities of the antioxidant enzymes did not
2 appear to be good indicators of O3 tolerance...."
3 Ozone exposure (70 ppb for 7 h/day for 14 to 42 days of exposure)10 caused an increase in
4 POD and a decline in SOD with no change in APX. No GSH was detected, but the
5 concentration of (ASC + DHA) was at 20 to 25 nmol/g-FW of extracellular fluid, compared with
6 2.4 to 3.0 mmol/g-FW of cell fluid. Glutathione within the cell was only 100 to 170 nmol/g-FW
7 of cell fluid. While these results are what one might expect for POD, the decline in SOD and
8 lack of change in APX are not what would be expected if protection was provided by SOD and
9 ascorbate. Yet as noted, because the rate of SOD reaction is many times higher than the rate of
10 O3 entry, there may be no pressure to increase the SOD level.
11 Some protection against visible injury (induced by 59 ppb daily mean O3 for 14 h/day for
12 7 days) was observed in genetically modified tobacco plants with excess chloroplast SOD11 (2 to
13 4 times higher), but less protection was observed in plants that had an excess of mitochondrial
14 SOD (8* higher) (Van Camp et al., 1994). In all lines, the conductance of the leaves dropped
15 about 50%, compared with the unmodified plants. There was a correlation with age of leaf (less
16 injury in younger leaves) that corresponded to that found in spruce trees in which the amount of
17 SOD declined in relation to the longer that needles were held on the tree (Polle et al., 1989).
18 A slightly different study, however, found no O3 protection with varied SOD within the needles
19 (Polle and Rennenberg, 1991). Interestingly, in maize, the synthesis of SOD (any form) was not
20 stimulated by O3 exposure (at 0.50, or 0.75 ppm for 8 h, variable times thereafter) but was by
21 exposure to 90% O3 (Matters and Scandalios, 1987). It may be that this high level of O3 does not
22 affect the SOD, or perhaps it stimulates and degrades the enzyme simultaneously.
23 The conclusions to be drawn from these results are not obvious. There seems to be SOD (a
24 Cu-Zn form) present in the apoplastic space of some plants, but it does not necessarily rise with
25 O3 exposure. Thus, either its concentration is sufficient to provide protection or it is not needed.
26 Over-expression of any SOD in other organelles may play a role, especially in the chloroplast
At a level of 70 ppb, the concentration of O3 in air was about 3.06 x KT6 mol/m3, which with the
conductance of 0.042 mol/m2 s, gives a flux rate of O3 of 1.27 x 10~8mole/m2 s. Converting the SOD rate of
23 units/g-FW into a SOD rate within the apoplastic space of 6.9 x 10~3 mol/m2 s, or about 500,00 times the
entry rate of O3.
The SOD enzymes were from Nicotiana plumbaginifolia with appropriate transit sequence for targeting
the correct organelle and expressed under control of cauliflower mosaic virus 35S promoter.
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1 (Cu-Zn or Fe forms); but it may be playing a secondary role due to other effects of O3 that
2 generate conditions in which light can overload the chloroplast and generate detrimental
3 circumstances, including the production of SO2 . In addition, SOD is developmentally expressed
4 in varied concentrations, so that long term exposure to O3 may alter each leafs developmental
5 age and, in turn, alter what level of SOD is observed. In any case, SOD does not seem to be the
6 primary antioxidant system to protect against O3.
7
8 Changes to the Plasmalemma
9 Reports of "peroxidation" generally occur within unicellular organisms subjected to very
10 high levels of O3 (in Chlorella [Frederick and Heath, 1970)] and in Euglena [Chevrier et al.,
11 1990)]). Heath (1987) determined that by the time biochemical events were altered and MDA
12 was produced in Chlorella, little permeability remained in the cells and most metabolic
13 pathways were greatly disrupted by the subsequent loss of substrates. In fact, MDA production
14 was concurrent with a high O3 uptake by the cell, indicating a complete opening of the cell and
15 associated with the concurrent inability to plate the cells on a glucose median (indicative of cell
16 death). Heath reached the conclusion that no one had proven that lipid oxidation was in any way
17 a part of the initial reactions of O3 with the cell, a conclusion confirmed by Mudd et al. (1997a).
18 An excellent review regarding the initial action induced by O3 within a plant (Kangasjarvi et al.,
19 1994) should be consulted. There is little data to show that lipids are attacked by O3 in any
20 living system that was not previously severely injured by O3. Most of the data suggesting lipid
21 attack by O3 has been demonstrated in plants subjected to O3 concentrations of 0.5 to 1.0 ppm for
22 several hours, during which gross wilting of the plant tissues usually occurs, suggesting extreme
23 water loss. It is not surprising that lipid and protein injury is observed under these conditions.
24 While those reports were useful in the 1960s and 1970s, they are not especially insightful now
25 when ambient levels of O3 are rarely above 0.2 ppm.
26
27 9.3.4 Wounding and Pathogen Attack
28 The decline of an enzyme is more difficult to measure than the rise of a new enzyme; an
29 increase from 0 to 2% may be within the precision of any assay, but a decrease from 100% to
30 98% is often masked by simple variation of the assay. Thus, measuring enzymes, which are in
31 great abundance in pre-fumigated tissue, is a risky operation. On the other hand, if O3 induces a
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1 general physiological change that has characteristics similar to other well-studied, stress-induced
2 changes, then O3 studies could "piggy-back" onto those studies to gain insight into the full scope
3 of metabolic alterations. It is now becoming clear that wounding and pathogen attack of plants
4 are similar to O3-induced changes in plants, and a reasonable hypothesis is that O3 must induce
5 one or more of the first steps seen in the wounding/pathogen-attack response.
6 Systemic acquired resistance (SAR) has been heavily investigated, and DNA probes have
7 existed for some time for a series of expressed genes (see Table 9-9). Several enzyme classes
8 are associated with O3 injury, including glucanases and peroxidases and others, such as the PR
9 proteins and chitinases. Thus, strong evidence exists from enzyme function and genetic material
10 that O3 induces an activation of a SAR-like response.
11 Mehdy (1994) described a model of how an elicitor produced by the pathogen attack
12 activates a G-protein, which opens the inward-flowing Ca2+ channel. The flow of Ca2+ into the
13 cytoplasm raises the internal level (at the jiM level) and activates a protein kinase that increases
14 the activity of the plasma membrane NAD(P)H oxidase and generates O2~. Superoxide
15 dismutase converts O2 into H2O2. Both O2 and H2O2 are responsible for the active oxygen
16 species response, which is believed to be a defense mechanism to kill the pathogen. In this
17 normal defensive reaction, a subsequent system induces either localized lipid peroxidation per se
18 or a membrane lipase to produce j asmonic acid or inositol triphosphate, which act as secondary
19 messages to activate the defense gene products.
20 Booker et al. (2004) found that G-proteins might be involved in the perception of O3 in the
21 extracellular region using A. thaliana G-protein null mutants. The activation of a passive inward
22 flow of Ca2+, e.g., by an O3-induced response, would serve the same function as activation of the
23 G-protein. Once the level of cytoplasmic Ca2+ rises, all else follows. It is suspected that
24 exposure of plants to O3 does just that, as Castillo and Heath (1990) demonstrated — the in vivo
25 fumigation of bean plants both inhibits the outward-directed ATP-requiring Ca2+ pump and
26 increases the passive permeability of Ca2+. It was thought that the calcium transporter system
27 has a sensitive sulfhydryl group which, if oxidized, would alter normal Ca2+ movements. In
28 addition, Dominy and Heath (1985) observed that the K+-activated ATPase (believed to be
29 involved in K+transport) was inactivated by in vivo exposure to O3 and that inactivation was
30 traced to a sensitivity sulfhydryl. Mudd et al. (1996) argued that several amino acids are very
31 sensitive to O3, including any with an exposed sulfhydryl. Thus, the O3-induced change in Ca2+
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Table 9-9. Gene Families and cDNA Clones Used as Probes for SAR (Ward et al., 1991)
Probe
Relevant Properties of Encoded Protein
Reference
PR-1
PR-2
PR-3
PR-4
PR-5
PR-1 basic
Basic class III
chitinase
Acidic class III
chitinase
PR-O'
Basic, glucanase
Basic chitinase
SAR 8.2
Acidic, extracellular; function unknown most
abundant PR protein in tobacco; >90% identical to
PR-lbandPR-lc
Acidic, extracellular b-1,3-glucanase, >90% identical
to PR-N and PR-O
Acidic, extracellular chitinase; also known as PR-O;
>90% identical to PR-P
Acidic, extracellular; unknown function; homologous
to C-terminal domain of Winl and Win2 of potato
Acidic, extracellular; homologous to thaumatin and
bifunctional amylase/proteinase inhibitor of maize;
also known as PR-R or PR-S
Basic isoform of acidic PR-1
Homologous to cucumber chitinase (Metraux et al.,
1989), structurally unrelated to PR-3
Extracellular; approximately 60% identical to basic
isoform
Acidic, extracellularb-l,3-glucanase; approximately
55% identical to PR-2 group
Vacuolar; approximately 55% identical to PR-2 group
and PR-O'
Payne etal. (1988b)
Ward etal. (1991)
Payne etal. (1990a)
Friedrichetal. (1991)
Payne etal. (1988a)
Payne etal. (1989)
Lawton et al. (manuscript in
preparation)
Lawton et al. (manuscript in
preparation)
Payne etal. (1990b)
Shinshietal. (1988)
Vacuolar; approximately 65% identical to PR-3 group Shinshi et al. (1987)
Unknown function; cloned by ± screen of cDNA
library from secondary leaves of TMV-infected plants
Acidic peroxidase Extracellular; lignin-forming
Alexander et al. (manuscript in
preparation)
Lagrimini and Rothstein (1987)
1
2
3
4
5
permeability may be the trigger to most, if not all, the wounding responses. However, the
difficult problem of proving that the cytoplasmic Ca2+ change is the first event in O3 injury
remains.
Some wound- and pathogen-induced genes that are activated or repressed in Arabidopsis
thalania are found with DNA arrays (Cheong et al., 2002). While these responses may not be
uniform for all plants, they suggest the possibility of wide-ranging gene changes that may occur
with a simple wound and that those changes are wide-ranging and diverse. As an example, these
responses are related to hormonal responses that are related to jasmonic acid, ET, and auxin
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1 pathways; signal transduction responses; and transcription factors for a variety of pathways.
2 The involvement of ET in wounding and pathogen attacks is discussed in Section 9.4.3.2.
3
4 9.3.4.1 Peroxidases
5 Increases in cytosolic and apoplastic peroxidase activity in response to O3 are often
6 observed, but the reasons and outcomes of these changes have yet to be fully explained.
7 Increased activity is frequently correlated with O3 injury. Dass and Weaver (1972) observed that
8 increases in peroxidase after O3 injury was similar to that observed for plant infection by a virus.
9 Tingey et al. (1975) observed a 35% decrease in peroxidase activity immediately following O3
10 exposure; however, within 24 to 48 h, activity had increased significantly and was above control
11 level and remained there throughout the remainder of the study. Dijak and Ormrod (1982) also
12 observed increases in peroxidase activity when two O3-sensitive and two O3-resistant varieties of
13 garden peas (Pisum sativum) were exposed to O3. Peroxidase activity was not related to cultivar
14 sensitivity nor to visible injury. Unfortunately, there are many peroxidases (Birecka et al., 1976)
15 therefore, any general increase is not specific. In ET-treated leaves, peroxidase reaction
16 products were found between the plasma membrane and the cell wall, suggesting that ET itself
17 could induce peroxidase activity (Abeles et al., 1989a,b; Birecka et al., 1976).
18 At the same time, others examined peroxidase reactions in general and found two types of
19 peroxidases (designated as acidic or anionic and basic or cationic, EC 1.11.1.7, but also listed as
20 EC 1.14.18.1). Many types of peroxidases are located in diverse organelles, and each seems to
21 be activated by different conditions (e.g., pH for anionic and cation types and substrates such as
22 guaiacol, syringaldazine, and ascorbate). Peroxidases belong to at least two groups, which
23 catalyze two separate reactions: (1) the reaction of H2O2 with ascorbate to form DHA, discussed
24 earlier (Thorn and Maretzki, 1985), which is regenerated by plasma membrane electron transport
25 using a dehydrogenase (Gross and Janse, 1977) that is now believed to be a malate/oxaloacetate
26 shuttle through the membrane coupled to a NAD(P)H-cytochrome-b-reductoxidase; and
27 (2) the reaction with coniferyl alcohol (from phenylalanine through phenylalanine ammonia
28 lyase) to form lignin within the wall. The anionic peroxidase thought to be involved with
29 lignification is within the cell wall (Buchanan et al., 2000; Taiz and Zeiger, 2002). Some basic
30 peroxidases are maintained within the cell, while some are external to the cell. After wounding
31 (Lagrimini and Rothstein, 1987; Gasper et al., 1985), some basic peroxidases can be activated by
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1 processes leading to the synthesis of stress ET (Yang and Hoffman, 1984) and/or by excess Ca2+
2 (Gasper et al., 1985). Elicitor treatment of plants change a series of peroxidases, some of which
3 are similar to those seen in O3-induced changes (see Table 9-9).
4 The formation of lignin is due to the phenylpropanoid metabolism (Buchanan et al., 2000).
5 Tyrosine and phenylalanine are converted to cinnamic and/>-coumaric acid, which are in turn
6 converted to/>-coumaryl, coniferyl, and sinapyl alcohols, and then into lignins. Hence, the
7 peroxidase activity is often measured by one of these substrates (Espelie et al., 1986; Gasper
8 et al., 1985). However, it is questionable whether apoplastic peroxidase activity is limiting for
9 lignification; laccases have a prominent role as well. Also, availability of monolignols is critical
10 for core lignin formation, and it is unclear whether levels of these metabolites change in
11 response to O3. Studies by Booker (Booker et al., 1991, 1996; Booker and Miller, 1998; Booker,
12 2000) indicated that O3 did not increase core lignin concentrations in foliage of loblolly pine,
13 soybean, or cotton; although levels of phenolic polymers and cell wall-bound phenolics were
14 elevated in soybean. Increased phenolic polymers appear to be lignin in acid-insoluble lignin
15 assays and may well be responsible, along with polyphenol oxidase, for the stippling injury
16 observed in O3-treated plants. Cell wall function implies the transport of peroxidase molecules
17 out of the cell and, most likely, the regulation of their activities within the wall space. These
18 extracellular peroxidases may be observed by vacuum infiltration of buffer into leaf air spaces
19 and subsequent centrifugation of the tissues to remove the buffer with the apoplastic enzymes,
20 that wash out (Castillo and Greppin, 1986). However exposure to O3 induces important changes
21 in the plant. For example, extracellular peroxidase activity in Sedum album leaves increased
22 nearly 3-fold over that in the control plants after 2-h exposure to 0.40 ppm (Castillo et al, 1984).
23 This O3-induced increase of extracellular peroxidase appears to be under the control of Ca2+
24 (Castillo et al., 1984; Heath and Castillo, 1987). Initially, no effect on the anionic activity as
25 measured with syringaldazine (specific electron donor for lignifying peroxidases) was observed,
26 yet 21 hours later, the anionic peroxidase activity was increased, whereas the cationic (ascorbate
27 measured) peroxidase activity was decreased in O3-treated plants. This suggests an immediate
28 response (ascorbate peroxidase activation) and a secondary response that activates the lignifying
29 peroxidase via gene activation.
30 The rapid response of cationic peroxidase after O3 exposure may not result from de novo
31 protein synthesis but from the secretion and direct activation by Ca2+ ions of enzyme molecules
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1 already present in the tissue. Cationic peroxidases might attack the peroxides and, in this
2 manner, act as a detoxifying agent with ascorbate as the substrate in the apoplasm. The effect of
3 Ca2+ upon peroxidase activity is stronger at low H2O2 concentrations (Penel, 1986). Thus, one
4 can imagine that, when the H2O2 concentration is low, this peroxidase activation would have a
5 greater in vivo importance. Furthermore, the secretion of cationic peroxidases into the free
6 spaces as a result of O3 treatment is accompanied by a simultaneous release of at least one of its
7 natural substrates (ascorbic acid); this cationic peroxidase exhibits a much higher affinity
8 towards ascorbate (up to 6-fold) than the anionic isozyme (Castillo and Greppin, 1986).
9
10 9.3.4.2 Jasmonic Acid and Salicylic Acid
11 Salicylic acid (SA) and jasmonic acid (JA) are considered to be regulators of the plant
12 defense response (Figure 9-13) (Buchanan et al., 2000). They tend to respond more slowly than
13 ET, causing widespread effects in the plant tissues. Both seem to be heavily involved in
14 responses of the plant to O3, once again linking the pathogen/wounding defense to O3-induced
15 injury; however, their roles are far from clear.
16 One of the lipoxidase isoforms is activated by pathogen infection (POTLX-3) within 6 h
17 and accumulates for a week (Kolomiets et al., 2000). This enzyme is the first stage of the JA
18 pathway which leads to 13-hydroperoxide linolenic acid (HPOT) which is converted either to
19 allene oxide through AOS or to C6 aldehydes through hydroperoxide lyase. These aldehydes act
20 as signaling agents via systemin (Sivasankar et al., 2000) or volatile odiferous compounds
21 (oxylipins) that have been implicated as antimicrobial toxins (Froehlich et al., 2001).
22 Interestingly, these compounds seem to target the chloroplast envelop where they interact with
23 its metabolism. As FIPOT and AOs are both implicated in plant defense and are activated by O3,
24 these interactions may be related to how chloroplast enzymes and their mRNAs are involved in
25 O3-induced injury.
26
27 9.3.4.3 Stress-Induced Alterations in Gene Expression
28 Early studies addressed the qualitative and quantitative effects of O3 on protein metabolism
29 (Harris and Bailey-Serres, 1994). Subsequent reports suggested that the physiologic and
30 metabolic consequences of exposure to O3 was, in part, mediated by increased gene expression.
31 A summary of the gene-linked changes in proteins induced by SAR may be seen in Table 9-9.
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Phospholipid-
Ethylene—
Salicylic Acid-
-> /AO Synthas^
0='
JasmonicAcid
(Howe ef a/., 2000)
COOH
Allene Oxide
if
Reductase
HOOC
13-Hydroperoxy-
. A • i
Imolenic Acid
(HPOT)
Traumatic Acid
(Croft etai, 1993)
COOH
Systemin
Figure 9-13. The pathway leading from phospholipids to jasmonic and traumatic acid.
The role of lipoxygenase and the production of a hydroperoxyl moiety from
the unsaturated fatty acid is clearly demonstrated. More importantly,
several of the enzymes within this pathway have been shown to be activated
by oxidative conditions including O3 exposure. The production of both of
these acid species could lead to a general global response of a whole plant to
the O3 exposure of a single leaf.
Source: Howe et al. (2000); Croft et al. (1993); Buchanan et al. (2000).
1 Of particular note are the productions of PR proteins, chitinase, glucanase, and acidic
2 peroxidases that appear to be common markers used in many O3 studies. A summary of varied
3 proteins as measured by changes in the mRNA in A. thalcmia induced by O3 exposure is shown
4 in Table 9-10. While studies on Arabidopsis thaliana required high concentrations of O3 to
5 produce a response, the levels reported in most of the studies did not induce visible injury. The
6 types of messages induced included glutathione S-transferase, PAL, ACC synthase, SOD, and
7 some PRs. Slower increases in messages are seen for other PR and SAR-senescence proteins.
8 Declines in messages were observed for varied chloroplast enzymes, including those for Rubisco
9 and chlorophyll binding proteins. A few new proteins were found — a casein kinase and three
10 plasma membrane proteins. It is interesting to note that few messages for "new" proteins were
11 generated by O3 exposure.
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fa
to
O
o
VO
«^j
o
H
6
o
0
H
O
o
0
o
H
W
Table 9-10. Proteins Altered by Ozone as Measured by Molecular Biological Techniques as mRNA Level or
Other Gene Activity Rather than Enzyme Activity
Exposure
150/300 ppb for 6 h
daily
300 ppb for 6 h daily
200/500/1 000 ppb
for2h
350 ppb for 1-6 h
300 ppb for 6 h
1 50 ppm for 6
h/8 and 14 days
160 ppb for 3-72 h
(a) 250 ppb for 8 h;
(b) 250 ppb for 2 h;
(c) 175 ppb for
8 h/4 days
Identified
proteins fast Slow increase
Physiological events increase response response
Leaf curling; GST, PAL Pxase, SOD
reduced growth
10 bands of 10RNA
Wilting (8 h); premature
senescence
Ethylene production; ACS-6
downward curvature;
water logging
Necrosis in NahG and Chi SOD. Chi GPX
Cvi-0 (accumulating SA), cytAPX, GST1
not in Col-0
Downward rolling of leaf; BCB, ERD 1 ,
early senescence SAG21
Early senescence GST1,VSP2; MT1
(a) little chlorosis or lesions; GSTApx, PAT1
(c) growth retardation CuZn-SOD
Examined, Unknown
Decline but no change proteins Reference
CAT, LOX1 Sharma
(1994)
AtOZIl » casein Sharma
kinasell (1995)
3 plasma Tokarska-
membrane Schlattner
proteins: 75-, 45-, (1997)
35-kDapeptides
ACS-1,-2, -4, -5 Vahala
(1998)
Cab mRNA, Rao (1999)
cyto SOD,
chlGR
Cab, rbcS Atgsr2, MT1 , Miller (1 999)
SAG 12, SAG 13,
SAG 19, SAG 20
CCH Mira (2002)
Fe-SODl GR, Conklin
cab,rbs (1995)
-------
to
o
o
Table 9-10 (cont'd). Proteins Altered by Ozone as Measured by Molecular Biological Techniques as mRNA Level or
Other Gene Activity Rather than Enzyme Activity
Exposure Physiological events
200 ppb for 24 h
Identified
proteins fast
increase response
PR-l,PR-2a,PR-5
AtEDSl,AtGSTl,
AtGST2
Slow increase
response Decline
PR-3b, PR-4
Examined, Unknown
but no change proteins
LOX2,AtOZIl,
PAL, Lhcb,
PAT1,HSP
Reference
Matsuyama
(2002)
250 ppb for 6 h
Lesion initiated on margin
and spread inward
rcdl, on
chromosome 1, Overmyer
single Mendelian (2000)
trait
to
Abbreviations used in Tables 5 and 6.
GST = Glutathionine synthase
PAL = Phenylalanine ligase
PR-1 =
Pxase = Peroxidase
CAT = Catalase
LOX1 = Lipoxygenase
ACS-6 = ACC synthatase
SOD = Superoxide dismutase
CuZn-SOD = cyto SOD
Fe-SODl = Chi SOD
cytAPX = Ascorbate peroxidase
Chi GPX = Glutathione peroxidase
Cab mRNA = Chlorophyll a/b binding protein
chl GR = Gluthatione reductase
BCB =
ERD1 = Ethylene response
SAG21 = Senescence
rbcS = Rubisco small subunit
MT1 = Mitochondria
H
6
o
o
H
O
o
HH
H
W
-------
1 The working hypothesis is that O3, which is not eliminated by antioxidants in the cell wall,
2 alters the properties of the plasma membrane. Specific polypeptides, indicative of these
3 antioxidants, are induced. If specific receptor molecules or channels on the membrane are
4 affected, the ionic balance within the cytoplasm is changed, leading to altered transcription or
5 translation of the genes controlling those and other types of polypeptides. Once this membrane
6 disruption occurs, the cell must mobilize repair systems to overcome the injury. Thus, carbon
7 and energy sources once destined for productivity, must be used in repair processes. Some of
8 these repairs are thought to result from the induction of specific genes. Photosynthesis is
9 inhibited by direct inhibition of some of the enzymes, through byproducts of O3 attack or by
10 altered ionic balance. At the very least, the decrease in photosynthesis is a result of an
11 O3-induced decrease in rbcS mRNA.
12
13 9.3.5 Primary Assimilation by Photosynthesis
14 9.3.5.1 Photooxidation: Light Reactions
15 Photooxidation refers to the oxidation of chlorophyll within the light reaction due to an
16 imbalance between light absorption and the CO2 use to produce carbohydrates. It was
17 discovered in the 1920s and studied under the concept of chlorophyll bleaching and photo
18 autooxidation (Asada, 1999; Rabinowitch, 1945). What generally occurs is that electron transfer
19 from H2O to NADPH declines, and a light reaction overload occurs. The slowdown of electron
20 transfer may also be due to inhibition of the dark reactions, through the poor use of small
21 molecular weight carbohydrates or a lowered amount of the fundamental substrate CO2.
22 To counteract these detrimental reactions, a series of "antioxidant" reactions exist, which
23 eliminate the buildup of oxidative intermediates.
24 A lowered CO2 level, which can be caused by stomatal closure (Heath, 1996), blocks the
25 use of reduced plastoquinone (PQH2) in Photosystem II through NADP reduction in
26 Photosystem I (Hankamer et al., 1997). The buildup of PQH2 reduces the amount of QA,
27 resulting in a buildup of P680+|Pheo~ species (the primary photoact). The inability to reduce this
28 radical leads to injury to the Dx protein (32 kDa) and its fragmentation into 23-, 16-, and 10-kDa
29 fragments (Hankamer et al., 1997). Ozone exposure of bean plants leads directly to the loss of
30 this Dj protein (Pino et al., 1995). The loss of Dj stimulates the production of new Dl (and its
31 mRNA). Also, the production of the oxidized form of P680 (P680+) is harmful to the plant because
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1 electron flow from water to P680+ is limited, generating a P680T (the triplet form of P680), which is
2 highly oxidizing and can lead to dangerous reactions. One form of protection is the use of
3 p-carotene to convert the triplet form back to its normal state; however, that reaction can lead to
4 the loss of P-carotene. Without the protection of P-carotene, oxygen with oxidized products to
5 produces singlet state of O2. This, in turn, can react with chlorophyll, leading to ring breakage
6 that, in essence, leads to chlorosis. These types of reactions do not seem to occur often, but
7 chlorosis is one form of visible injury, and loss of P-carotene has been reported. Farage et al.
8 (1991) and Farage and Long (1999) studied these reactions in wheat and concluded that
9 alterations to the dark reactions were much more common.
10
11 9.3.6 Alteration of Rubisco by Ozone: Dark Reactions
12 A large body of literature shows that O3 exposure induces a decline in Rubisco (Pell et al.,
13 1997). Treatment of a variety of plants with O3 at near ambient levels results in a loss of
14 Rubisco and of the mRNA coding for both subunits of Rubisco (rbcS, small and rbcL, large).
15 Because Rubisco plays such an important role in the production of carbohydrates (Figure 9-14),
16 any loss may have severe consequences for the plant's productivity.
17 The study by Noormets et al. (2001) used an exposure system of plants in a FACE
18 exposure facility in which areas of ambient (daytime 360 ppm) and ambient with added CO2
19 (560 ppm), with added O3 (97.8 ppb), and with added CO2 and O3, were used. Two clones of
20 aspen (O3-tolerant and -sensitive) were used in all four cases. The study recognized that leaf age
21 was important in the variation and tried to control for it by grouping the data by leaf plastochron
22 index12. These data, taken over many days using the LiCOR 6400, confirmed earlier
23 observations that O3 has the greatest effect on older leaves after causing a decline in assimilation
24 and in conductance. They also confirmed that the internal CO2 (calculated for within the leaf) is
25 not affected by O3 exposure. Higher levels of CO2 increased the assimilation and lowered the
26 conductance, maintaining the internal to external CO2 ratio identical to that found with the
27 ambient CO2 level, corresponding to the theory of Farquhar et al. (1980). The level of Rubisco
28 was not measured as frequently as the assimilation and conductance, but the significant (5%)
29 increase in Rubisco was only observed in older leaves for both clones. More to the point was
12The plastochron index for the leaf measures the leaf age by degree of expansion rather than simply by
chronological time.
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{precursors}
{precursors}r
DNA —*> {rbcJi
{products}r
{products}
{Triose-P}
Figure 9-14. The production of Rubisco and its Calvin Cycle pathway reactions. Two
peptides are used to build Rubisco: rbcS, the small subunit produced by
DNA within the nucleus; and rbcL, the large subunit produced by DNA
within the chloroplast itself. Clearly both polypeptides must be closely
regulated to produce the enzyme in a coherent manner. Furthermore,
at least five isoforms of DNA can produce rbcS, each of which is regulated
by a different promotor region.
1 that the stomatal limitation13 was not altered by O3 exposure, with or without excess CO2. It is
2 critical to point out that mesophyll conductance is directly linked to internal CO2 level14. So if
3 Ci/Co is constant and gs declines, then gm must likewise decline. If, as it is argued, Rubisco
4 levels are constant or at least increasing, then a regeneration of RuBP must be the cause of the
5 decline in gm. Farquhar et al. (1980 were more concerned with high levels of CO2 and had little
13The limitation was defined as the ratio of stomatal resistance to the total resistance (which included the
operating point of the assimilation (A) verses internal CO2 concentration (Q) curve and the resistance of the
boundary layer). The operating point of the curve was defined as the internal CO2 level (which is calculated by the
conductance and assimilation). The resistance of this operating point was calculated as the cotangent of the slope to
the operating point. Unfortunately, the slope is not a dimensionless parameter but is rather moles of air per area of
leaf ~s of time and, thus, it is unclear whether the slope changes with added CO2 and O3.
14Respiration is generally small at saturating A and often is ignored. By transforming {A = gs (C0 - Q)}
into {A = gs C0 - gj,, C0 = (gs - g,,,) Co} where gm= gs (Q/C0) or the mesophyll conductance in earlier literature.
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1 to say about O3 exposure. One thing is certain however, the use of clones using different aged
2 leaves is a preferred method for this type of investigation.
3 The level of carbohydrate within the cell has an effect upon the amount of mRNA for
4 Rubisco (rbcS). Experiments by Krapp et al. (1993) indicated that a decline in carbohydraft
5 levels is probably due to the increased production of control metabolites, such as fructose
6 2,6-bisphosphate, which can shut down important sugar production pathways. This report also
7 leads to a measure of half-time for the decline in rbcS of about 2 days15 when 50 mM glucose is
8 added to a cell suspension of Chenopodium. Also, the carbohydrate level was increased by cold
9 girdling the petioles of tobacco and potato using intact plants. The levels of carbohydrate nearly
10 doubled in 5 days and the level ofrbcS declined rapidly (reaching 25% after 12 h). A decline in
11 Rubisco followed, but more slowly (with an estimated half-time of about 108 h after a lag of at
12 least 12 hours). This, of course, is expected; the level of the enzyme would decline slowly with
13 a lag after a loss of the message.
14 A better estimation of the half-life ofrbcS can be found in the Jiang et al. (1993) study of
15 the destabilization of the message by an antisense message. The wild type rbcS in tobacco had a
16 half-life of about 5 h compared to that in the mutant with the antisense. It was argued that the
17 antisense message increased the degradation of the normal rbcS. The estimated half-life ofrbcS
18 under O3 fumigation is about 1 hour (Pell et al., 1994). Although comparisons of these diverse
19 systems can not be easily made, the normal half-life ofrbcS may be closer to 5 to 10 h; and O3
20 fumigation does not simply stop the transcription of DNA, but rather it alters the rate of
21 degradation, either independently of, or simultaneously with, transcription.
22 Williams et al. (1994) developed a correlation between the levels of ABA after water stress
23 in Arabidosis thaliana leaves and the loss ofrbcS. Although their data were not quantitative, the
24 level of ABA had a half-time rise of about 1 to 2 h and the level ofrbcS had a half-life decline of
25 about 2 to 4 hours. Their work suggests that drought stress may alter the CO2 metabolism by
26 changing enzyme relationships much more than by merely closing the stomata. If an ABA rise
27 is lowering rbcS, rbcS may not be a good marker of O3 fumigation except under highly
28 controlled conditions.
29
15The amount of Rubisco drops from an initial 0.12 to a final amount of 0.04 umole/g-FW s in 6 days.
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1 9.3.7 Carbohydrate Transformations and Allocation
2 The question of whether translocation of the sugars out of the leaf is inhibited by O3
3 exposure arises because productivity is often dramatically inhibited by O3 fumigation. Though
4 nearly 35 years have past since Dugger and Ting (1970) investigated the question of sugar
5 transport within the leaf, the question has since been little studied. Translocation (Cooley and
6 Manning, 1987) appears to be inhibited, because root functions are impaired by O3 exposure.
7 Many observed events suggest that while carbon assimilation within the leaf declines,
8 translocation of carbon is inhibited even more so, because plant growth points are inhibited and
9 root/shoot ratios are altered (Dugger and Ting, 1970; Gerant et al., 1996; Tjoelker et al., 1995).
10 Many of the experiments with O3 fumigation indicate that O3 exposure decreases the net
11 growth or dry mass of the plant, but the mechanism is poorly understood. Generally the
12 decrease in assimilation is much less than the decrease in growth, but not always. Under many
13 conditions, the stomata will close partially, decreasing assimilation by a smaller factor. Only a
14 long exposure, or high levels of exposure for a short time, generate enough decline in Rubisco to
15 make the growth of the plant problematical. No convincing argument has linked the decrease in
16 growth with a small decline in assimilation, either by a conductance- or Rubisco-limitation.
17 Measures of assimilation with crops are frequently done on upper canopy leaves, which are the
18 last leaves to exhibit O3 injury, while leaves deeper in the canopy exhibit injury and early
19 senescence. Crop root growth must be sensitive to these and other O3 effects, because root
20 biomass is often suppressed early by elevated O3.
21 Volin et al. (1998) found O3 exposures statistically decreased leaf area ratio, specific leaf
22 area, leaf weight ratio, and root weight ratio in Populus tremuloides and two C3 grasses
23 (Agropyron smithii and Koeleria cristatd) but not in Quercus rubra and in the C4 grasses
24 Bouteloua curtipendula and Schizachyrium scoparium. There was no statistically significant
25 change in any species in leaf conductance (4% level decline in K. cristatd) nor in assimilation
26 (although there was a decline in assimilation at the 6% level for P. tremuloides and a decline at
27 the 1% level in B. curtipendula). They also reported a correlation between growth decline and
28 decreased stomatal conductance among all species.
29 Trade-offs are made by plants. Birch grown in highly fertilized conditions exhibited a
30 greater leaf turnover when exposed to O3, in that leaves not only formed faster but abscised
31 faster, presumably due to early senescence; whereas birch grown under poorer fertilized
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1 conditions retained their leaves longer and had a greater respiration rate within those leaves
2 (Maurer and Matyssek, 1997). Again, one must be careful in comparing short-term versus long-
3 term exposures. Grulke et al. (2001) observed that maximum concentrations of carbohydrates in
4 1-year old needles that had not abscissed due to early senescence declined when subjected to
5 year-long exposures along an increasing pollution gradient. Furthermore, the monosaccharide
6 concentrations in fine roots (along with starch) were largely decreased, suggesting that needle
7 sugars were limiting, leading to root sugar limitations. However, determination of the total
8 productivity and detailed balance of carbohydrate was impossible, because these were older,
9 larger trees and the data were taken over a full growing season. For a shorter-term exposure of
10 9 days, Smeulders et al. (1995) observed that O3 appeared to increase the retention of labeled
11 photosynthates within the needle, and, at higher exposures (400 versus 200 or 0 mg/m3), the total
12 starch within the needle decreased, suggesting that less carbohydrate was produced within the
13 cell or perhaps that it was in compounds not measured.
14 Studies with Pima cotton, aspen (Populus spp.) and bean seedlings (Phaseolus vulgaris)
15 indicate that acute O3 exposures inhibit export of the current assimilate that provides
16 carbohydrates to the roots from source leaves of cotton as well as recent assimilate from the
17 older leaves of aspen and bean (Grantz and Yang, 2000). Grantz and Yuan (2000) attempted to
18 distinguish between potential mechanisms of O3 phytotoxicity operating at the level of the whole
19 plant. Four hypothesis were tested by fumigating cotton: (1) O3 exposure reduces leaf pools of
20 soluble sugars; (2) pruning leaf area and reducing source strength to match that of O3-treated
21 plants reproduces O3 effects; (3) pruning lower leaf area more closely reproduces O3 effects than
22 pruning the upper leaf area; and (4) manipulating plant age and, thereby, plant size to match
23 O3-treated plants reproduces O3 effects. All were shown to be incorrect. Under each of the
24 above conditions, Grantz and Yang (2000) reduced the amount of foliage to match that caused
25 by O3 injury. While the treatments reduced the biomass and leaf area, they did not alter biomass
26 allocation nor root function. They concluded that a simple loss of foliage does not induce the
27 changes in translocation to the roots to the same extent as does O3 injury.
28 This finding by Grantz and Yang (2000) is important in that it suggests that O3 can trigger
29 a plant-wide response that may be linked to alterations in signal transduction and the generation
30 of whole plant signals. Stitt (1996) suggested that"... allocation is regulated by long-distance
31 signals that act to influence growth of selected sinks and to modify the delivery of resources to
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1 these sinks in parallel." Cooley and Manning (1987), citing Me Laughlin and McConathy
2 (1983), suggested three possible ways that O3 fumigation might alter translocation:
3 (1) malfunction of the phloem loading process, (2) increased allocation to leaf injury repair, and
4 (3) an altered balance between the leaf and sinks caused by reduced carbon fixation and a greater
5 demand for assimilate in the leaf.
6 Ethylene has been shown to reverse this sugar inhibition of development and to be
7 antagonistic to the ABA effect (Finkelstein and Gibson, 2002). However, these effects depend
8 greatly upon the developmental stage of the plant. Thus, the balance of the effectors (sugars,
9 ABA, and ET) may interact to generate the variation observed in the O3-induced productivity
10 decline. For example, O3 fumigation can induce a shift in the carbon transfer between roots and
11 shoot; and this shift can be amplified by mild drought (Gerant et al., 1996). Furthermore a
12 regulation of source-sink relations with the defense responses induced by elicitors was observed
13 by wounding the leaves of Chenopodium rubrum. Ethylene appears to be able to repress the
14 expression of extracellular invertase, which is critical for control and down-loading of sucrose
15 derived from the translocational stream (Roitsch, 1999; Lindow et al., 1996). In addition, the
16 development of Arabidopsis at high concentrations of glucose or sucrose is arrested by
17 increasing the ABA level (Coruzzi and Zhou, 2001).
18 Clearly more work is needed on the interactions between assimilation, translocation, and
19 source/sink relations with O3 exposure. In these interactions, one must be aware of the
20 developmental age of the plants and their phytohormonal status.
21
22 9.3.7.1 Lipid Synthesis
23 Heath (1984) summarized several early reports of O3-exposure induced lipid alterations.
24 Most concerned the production of MDA as a measure of lipid oxidation as well as the loss of
25 unsaturated fatty acids. However, a series of experiments by Sakaki and coworkers concentrated
26 on one type of fumigation system and one metabolic pathway. This literature provides the best,
27 most complete story with regard to lipid metabolism and O3 fumigation and indicators that O3
28 injures cellular membrane systems via lipid destruction.
29 Sakaki and coworkers used spinach, which is a sensitive plant but which has not been
30 much evaluated with respect to O3 fumigation. While the O3 level was high (0.5 ppm), enough
31 work has been done to be able to "tease apart" what is happening. The first paper showed that
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1 chlorophyll bleaching does not begin until the plants have been exposed to O3 for over 10 h,
2 whereas some MDA production begins with as little as 6 h exposure (Sakaki et al., 1983).
3 Consistent production of MDA, indicative of gross disruptions, occurred only after 8 h exposure
4 (Sakaki et al., 1985), within the timescale when chlorophyll and carotenoid levels began to
5 decline. Concurrently, the total fatty acid (FA) level decreased from -481 to 358 nmol/cm2 as
6 the MDA level increased from 0.6 to 2.4 nmol/cm2, indicating FA peroxidation (Sakati et al.,
7 1985).
8 Sakaki et al. (1983) also studied development of changes by cutting disks from exposed
9 leaves and floating them on water solutions for varied time periods (up to 24 h). This permitted
10 feeding experiments to be done easily, whereas the cutting gives rise to an additional wound
11 response and eliminates metabolite movement to and from other portions of the plant. The
12 floating experiments indicated that, after exposure, scavengers of singlet oxygen (JO2), such as
13 D2O, and of hydroxyl radicals, such as benzoate and formate, have no effect on development of
14 the MDA response after 8 h of in vivo fumigation, while scavengers of (O2 ), such as tiron and
15 ascorbate, lowered the amount of MDA formed. By measuring metabolites immediately after
16 cessation of fumigation, they were able to show that ascorbate loss began with the onset of
17 fumigation, as did SOD loss. A lag time associated with the production of DHA suggested that
18 the reaction of ascorbate with fumigation did not immediately produce the oxidation product.
19 The first 4 h of exposure yielded 30 nmole/cm2 of ascorbate loss with 5 nmole/cm2 of DHA
20 production, whereas the second 4 h of exposure yielded 20 nmole/cm2 of ascorbate loss with
21 20 nmole/cm2 of DHA production.
22 Nouchi and Toyama (1988) exposed Japanese morning glory (Ipomea nil) and kidney bean
23 (Phaseolus vulgaris) to 0.15 ppm O3 for 8 h. Under these conditions, little visible injury was
24 found with up to 4 h exposures, while injury increased by -50% after 8 h of exposure. Morning
25 glory produced more MDA than kidney bean, which produced the same as the zero-time control.
26 Morning glory also demonstrated a slight (5%) drop in MDGD (monogalactosyldiacylcerol),
27 with increases in PC (phosphatidylcholine), PG (phosohatidylglycerol), PI (phosphatidylinositol),
28 and PE (phosphatidyelthanolamine) after 4 h. Twenty-four hours later, the drop in MGDG
29 (mongalactosyldiacyglyerol) was much larger and was thought to be related to an inhibition of
30 UDP-galactose galactotransferase due to a rise in free fatty acids (FFAs) in the chloroplast. Note
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1 that the two distinct timescales involved in O3 fumigation, immediately post-fumigation and a
2 day or so later, allows for comparison after plant metabolism responds to the fumigation event.
3 The pathway for the formation of MGDG and DGDG (digalactosyldiacylglcerol) is located
4 on the chloroplast envelope. Diacylglycerol (DG) arrives from either the endomembrane system
5 or the stroma and the enzyme UDP-Galactose: 1,2-diacylglycerol galactosyltransferase (UDGT)
6 forms MGDG with galactose from UDP-galactose. Sakaki et al. (1990) suggested that the
7 O3-induced inhibition of UDGT was due to a release of FFAs from within the chloroplast. These
8 FFAs are inhibitory to UDGT, but not to GGGT, which is stimulated by high concentrations of
9 Mg2+ (Sakaki et al., 1990). The Sakaki et al. (1990) data indicate that the measured activities of
10 both enzymes isolated after fumigation has indicated that in vivo are not affected by O3
11 fumigation in vivo. Both enzymes have sensitivity sulfhydryls, and both are located on the
12 envelope. Ozone, if it reaches those sulfhydryls, should inhibit these enzymes; yet inhibition
13 was not seen.
14 It has been thought for years that tocopherols functioned as antioxidants in biological
15 systems (Tappel, 1972). Hausladen et al. (1990) examined the role of antioxidants in red spruce
16 by following seasonal changes. They fit the level of tocopherol within the needles (/g-FW) to
17 the time of the year and found little change (fit as level = A + Bt + Ct2). From this empirical fit,
18 they found that the constant A was lower with higher levels of O3. The seasonal variation
19 coefficients, B and C, were also lower, suggesting year long lows tocopherol levels. Variation
20 with the season is not particularly surprising, given that phytochrome action may be linked to
21 tocopherol biosynthesis (Lichtenthaler, 1977). Hansladen et al. (1990) reported a significant
22 (p < 0.05) trend in the difference between the high and low level of treatment; although there
23 was no discussion of why it occurred or what it meant in relation to metabolism. Their major
24 conclusion was that the antioxidant changes due to O3 exposure may decrease cold hardiness.
25 Sterols, believed to act as membrane stabilizers, have been investigated by several groups
26 with mixed results. Tomlinson and Rich (1971), who exposed common bean at 0.25 ppm for
27 3 h, and Grunwald and Endress (1985), who exposed soybean at 0.07 ppm for 6 h for 48 days,
28 reported an increase in free sterols and a decline in esterified sterols. However, Trevanthen et al.
29 (1979) exposed tobacco at 0.3 ppm for 6 h and reported opposite results. None of these
30 investigators believed that O3 had attacked the sterols directly, instead, they believed that these
31 changes involved metabolism and membrane stability. If O3 induced a metabolic shift that
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1
2
3
4
5
6
7
8
9
10
11
12
13
disturbed the polar lipid to sterol balance, membrane reactions to other stressors, such as cold
tolerance, would certainly also be affected, perhaps detrimentally.
9.3.8 Role of Age and Size Influencing Response to Ozone
Clearly many changes occur with O3 exposure can be observed within hours, or perhaps
days, of the exposure. This document has argued that many of those events are connected with
wounding and elicitor-induced changes in gene expression, but those are relatively fast acting
changes (a timescale of tens of hours). Two other effects due to O3 take longer to occur and tend
to become most obvious under long periods of low-O3 concentrations. These have been linked
to senescence or some other physiological response very closely linked to senescence. These
two responses, separated by a time sequence, are shown diagrammatically in Figure 9-15.
03
M— Ascorbate •<—
"First Event"
{Membrane Alteration} U-
i ?
ROS
I
Hypersensitivity Reactions
Accelerated
Senescence
Figure 9-15. Linkage of senescence with hypersensitivity reactions and first event of O3
attack of plants.
1 The understanding of how O3 affects long-term growth and resistance to other biotic and
2 abiotic insults in long-lived trees is unclear. Often, the conditions to which a tree is subjected to
3 in one year will affect the response of that tree in the next year. This has been called "memory
4 effect", although the term "carry-over" is preferred. In other words, a condition in an earlier
5 year sets the stage for a reaction in the next year; thereby giving a "cause-effect" scenario.
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1 In perennial plant species, growth affected by a reduction in storage carbohydrates may
2 result in the limitation of growth the following year (carry-over effects). Carry-over effects have
3 been documented in the growth of tree seedlings (Hogsett et al., 1989; Sasek et al., 1991; Temple
4 et al., 1993; U.S. Environmental Protection Agency, 1996) and in roots (Andersen et al., 1991;
5 U.S. Environmental Protection Agency, 1996). Accumulation of the carry-over effects over time
6 will affect survival and reproduction. Data on the cumulative effects of multiple years of O3
7 exposures have been, for the most part, the result of 2- to 3-year seedling studies. The difficulty
8 of experimentally exposing large trees to O3 has lead to the tacit assumption that seedling
9 response to O3 is a good predictor of large-tree response to O3 (U.S. Environmental Protection
10 Agency, 1996).
11 The carry-over effects of O3 exposures as observed in tree seedlings cited above by Hogsett
12 et al. (1989) have been termed "memory effects" by Langebartels et al. (1997) and proposed by
13 Schmieden and Wild (1995) to explain the sensitivity of spruce seedlings to frost in the winter
14 after having been exposed to O3 during the previous summer. Norway spruce exposed to 80 ppb
15 for a whole growing season, demonstrated visible injury symptoms the following year when the
16 new needle flush appeared (Langebartels et al., 1997). Additional studies using Norway spruce
17 and Scots pine seedlings have showed similarly delayed responses following O3 exposures.
18 Carry-over symptoms were noted to develop at different times of the year, depending on the
19 species of seedling exposed: in early spring for Norway spruce, and in early autumn for Scots
20 pine (Lange et al., 1989). Visible effects of O3 exposures on spruce and pine may develop after a
21 substantial delay during the "sensitive" periods of the year when chlorophyll and needle loss
22 normally occur. Norway spruce and Scots pine differ in their sensitive periods because of the
23 different needle classes normally remaining on the tree (Langebartels et al., (1997).
24 Nutrient status of the tree during the over-wintering phase of its life (Schmieden and Wild,
25 1995) and chronic exposure to ambient O3 (less severe with fewer peaks of very high levels)
26 induce (1) mineral nutrient deficiency; (2) alterations of normal metabolism, including
27 translocation and allocation of carbohydrates and probably nitrogen; and (3) disturbance of
28 normal transpiration and diurnal cycling, leading to water stress. This condition, termed
29 "Montane yellowing", appears to be related to nutrient deficiencies rather than senescence
30 (although early loss of leaves and needles does occur). While generalized low nutrient
31 concentrations may not occur within the foliage, localized deficiencies may. However, they are
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1 hard to observe or prove without a great deal of work involving all portions of a tree and without
2 a general hypothesis of what is occurring.
3
4 9.3.9 Summary
5 As the understanding of wounding responses of plants and more genome details and varied
6 plant mutants become available, the cellular and physiological responses of plants to O3
7 exposures are slowly becoming clearer. However, more studies are needed on a larger variety of
8 species. Nevertheless, several key findings and conclusions can be highlighted:
9 (1) The entrance of O3 into the leaf through the stomata remains the critical step in O3
sensitivity. Not only does O3 modify the opening of the stomata, usually closing it
partially, but O3 also appears to alter the response of stomata to other stressful
situations, including a lowering of water potential and ABA responses. The
concentration of O3 within the leaf is not the same as the external concentration due
to reactions within the leaf, but it is not "zero".
10 (2) The initial reactions of O3 within the leaf are still unclear, but the involvement of
H2O2 is clearly indicated. The detection of possible products by EPR spectroscopy
has progressed, but has not reached the point where any products can be identified.
Nonetheless, reaction of O3 (or its product) with ascorbate and possibly other
antioxidants present in the apoplastic space of the mesophyll cells is clear and
serves to lower the amount of O3 or product available to alter the plasma membrane
of the cells.
11 (3) The initial sites of membrane reactions seem to involve transport properties and,
possibly, the external signal transducer molecules. The alteration and mechanism of
the alteration of the varied carriers of K+and Ca2+ is far from clear, but it would
seem that one of the primary triggers of O3-induced cell responses is a change in
internal Ca2+ levels.
12 (4) The primary set of metabolic reactions that O3 triggers now clearly includes those
typical of "wounding" responses generated by cutting of the leaf or by
pathogen/insect attack. Again, this seems to be due to a rise in cytoplasmic Ca2+
levels. Ethylene release and alteration of peroxidases and PAL activities, as well as
activation of many wound-derived genes, seem to be linked to some of the primary
reactions.
13 (5) The alteration of normal metabolism due to wounding has effects outside of the
cytoplasm. What effects are due to the "spreading of the problem" to other cellular
organelles is less clear. One of the secondary reactions is linked to an activation of
a senescence response. The loss of Rubisco and its messenger RNA is linked to an
early senescence or a speeding up of normal development leading to senescence.
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The loss of photosynthetic capacity is linked to the lowered productivity of plants,
and problems with efficient translocation are indicated, although photosynthesis and
translocation still occur at a reasonable rate. The loss of productivity is not yet
clearly explained.
14 It is important to note that the dramatic strides in understanding the genetic makeup of
15 plants, gene control, and signal transduction/control over the last few years will likely accelerate
16 in the future. That understanding will translate into better models of the hypotheses listed above
17 and more detailed schemes of how O3 alters much of basic plant metabolism. Thus, while
18 understanding of how O3 interacts with the plant at a cellular level has dramatically improved,
19 the translation of those mechanisms into how O3 is involved with altered cell metabolism, with
20 whole plant productivity, and with other physiological facts remained to be more fully
21 elucidated.
22
23
24 9.4 MODIFICATION OF FUNCTIONAL AND GROWTH RESPONSES
25 9.4.1 Introduction
26 The responses of plants to any air pollutant may be significantly influenced by a wide
27 range of biological, chemical and physical factors. A plant's genetic make-up is an important
28 inherent biological determinant of its response, but response can also be modified by other
29 biological agents such as disease-causing organisms, insects and other pests, and by other higher
30 plant species with which it may be competing for resources. Chemical factors that may
31 influence response range from mineral nutrients obtained from the soil to other air pollutants and
32 agricultural chemicals. Physical factors that may influence response include light, temperature
33 and the availability of moisture, which are components of climate and climate change.
34 Some environmental factors are capable of being controlled, to some degree, by man,
35 while others are not. The biological factors (pests, diseases, symbioses and competition) are
36 partly controllable in agriculture but much less so (if at all) in natural ecosystems. It is possible
37 to control agricultural soil fertility and the use of agricultural chemicals, as well as to exercise
38 some control over the supply of water and airborne chemical factors. In contrast, the physical
39 factors, i.e., light and temperature, are uncontrolled in the field even though they may be
40 controllable in specialized situations such as greenhouses or shade houses.
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1 The impacts of these various factors on plant response to O3 and other oxidants were
2 extensively reviewed in the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996).
3 It was noted in that document that, since any combination of these factors may come into play at
4 some time in a plant's life history, "response will be dictated by the plant's present and past
5 environmental milieu, which also includes the temporal pattern of exposure and the plant's stage
6 of development." That document also stressed that both the impact of environmental factors on
7 response to oxidants and the corollary effects of oxidants on responses to environmental factors
8 have to be considered in determining the impact of oxidants on vegetation in the field. The
9 variability observed in plant responses to defined exposures to O3, particularly under field
10 conditions, is a consequence of the influences of genetics and the range of environmental
11 variables.
12 In view of the large number of factors to be considered and given that the purpose of this
13 document is to support the review of the O3 NAAQS, including standards to protect vegetation,
14 this section focuses mainly on situations in which there is clear evidence that environmental
15 factors truly interact with oxidant effects, i.e., they magnify or diminish the impact of O3 and are
16 not merely additive to it. Conversely, it will cover situations where O3 acts synergistically or
17 antagonistically, but not additively, with effects induced by other factors. It will also emphasize
18 those interactions as a result of which overall plant growth and development, and yield are
19 adversely affected, rather than the details of interactions at the mechanistic level, unless the latter
20 are deemed to be essential to an understanding of larger scale effects.
21 To facilitate cross-reference, the present document uses essentially the same subsections as
22 in the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996). Although light and
23 temperature are components of climate, they are initially reviewed as individual physical factors,
24 even though temperature effects are revisited to some extent in the discussion of interactions
25 with climate change.
26 Few studies reported since the 1996 document have systematically investigated
27 quantitative responses to O3 concurrently with other variables. Although the 1996 document
28 cited almost 300 references pertaining to environmental interactions, and the present review cites
29 more than 350 new references, the bulk of the recently published work has continued to be
30 determined by the specific and frequently narrowly focused interests of individual researchers or
31 groups. Hence, the new findings are scattered and far from uniformly distributed among the
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1 various sub-topics. In some instances there has been little or no research published that adds to
2 our understanding since the 1996 document. In such cases, the present review is therefore
3 restricted to summarizing the understanding that was current in 1996.
4 A few reviews have appeared since the early 1990s dealing with various environmental
5 interactions, and these are cited in relevant sections below. More general recent reviews are
6 those of Wellburn (1994); multi-authored volumes edited by Alscher and Wellburn (1994),
7 Yunus and Iqbal (1996), De Kok and Stulen (1998), and Bell and Treshow (2002); and reports
8 by the United Nations Environment Programme (UNEP) (1999) and the Intergovernmental
9 Panel on Climate Change (IPCC) (2001). Several biotic and abiotic interactions involving forest
10 trees are discussed in the review by Johnson et al. (1996b).
11 Although many reports have provided quantitative information on interactive effects, in
12 most cases the information is only descriptive of a specific situation involving only two or three
13 levels of a variable. While this may be adequate to provide statistical information about the
14 existence of interactions with environmental factors, it does not permit the development of
15 response surfaces or models to show the form that any influence of such factors might take on
16 O3 exposure-response relationships or how O3 might quantitatively influence responses to the
17 factors in question. This, together with the fragmented information available on the effects of
18 most factors, has contributed to the relative lack of development of simulation models of
19 oxidant-environmental factor interactions. Yet, as noted by Taylor et al. (1994), the large
20 number of variables constrains the assessment of pollution effects by experimentation alone.
21 The only alternative is to use mathematical models to attempt to predict the outcome of different
22 O3 and environmental factor scenarios, building up their complexity in stages. The few models
23 thus far used to investigate O3 stress have been adapted from existing process models of crop or
24 tree growth which include limited numbers of physical or chemical variables (such as
25 temperature, soil water stress, or nutrient deficiency). Taylor et al. (1994) provide a listing of
26 several simulation models developed for trees at the individual, stand, and regional levels, and
27 these and many other models have been critically reviewed by Kickert and Krupa (1991) and
28 Kickert et al. (1999). However, regardless of whether such models are descriptive/empirical or
29 process/mechanistic, their outputs will always be associated with varying degrees of uncertainty
30 and require validation against observable responses wherever possible. Kickert et al. (1999) also
31 point out that very few of the models that have been described provide risk assessments that
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1 address likelihood, in contrast to consequence assessments that address the magnitudes of
2 effects. Thus, even though capable simulation models of plant response to O3 involving
3 complex mixes of many biological, physical and chemical factors may be out of reach at the
4 present time, the use of newer mathematical approaches such as artificial neural networks
5 (ANNs) has enabled insightful analyses to be performed in several field studies involving
6 numerous micrometeorological and other environmental variables (e.g., Balls et al., 1996;
7 Mills et al., 2000).
8 Because the ensuing subsections deal with studies of O3 interactions involving an
9 extremely wide array of biological, physical, and chemical factors in the plant's environment,
10 it is inevitable that many different exposure facilities and regimes have been used in these
11 studies. To provide specific information regarding the O3 exposure concentrations, profiles,
12 hours and days of exposure (as well as the types of systems and facilities used for the exposures)
13 would add a wealth of detail that would do little to assist our understanding of the roles of
14 environment factors in modifying the impact of O3 on vegetation or to facilitate our ability to
15 estimate the magnitudes of any such modifications. Thus, only experiments in which the
16 exposure levels and regimes were within the bounds of ambient experience in North America
17 are discussed in the ensuing subsections, regardless of the type of exposure profile used. The
18 cutoffs used have been: -200 ppb for peak hourly concentrations or for short-term exposures;
19 -100 ppb for daytime means involving prolonged exposures for several hours; or a doubling of
20 ambient levels, in cases in which enriched exposure levels were a function of ambient levels.
21 Actual details of the exposure regimes and conditions can, of course, be obtained from the
22 original references but are only stated here when any distinction needs to be made between the
23 effects of different exposure levels. Hence, it should be understood that ensuing statements such
24 as "... it was found that O3 caused ..." should always be read as "... it was found that exposures to
25 O3 (within the range of those that have been measured in ambient air) caused ..."
26
27 9.4.2 Genetics
28 The response of individual plants to O3is affected by several factors including the
29 environment in which it is growing, competition with neighbor plants, ontogeny, and genetics.
30 This section examines the role of genetics in plant response to O3. In addition, major knowledge
31 gaps in the understanding of genetic aspects of O3 response are pointed out.
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1 It is well known that species vary greatly in their responsiveness to O3 (Environmental
2 Protection Agency, 1996). This again has been recently demonstrated for grassland species
3 (Warwick and Taylor, 1995; Pleijel and Danielson, 1997; Grime et al., 1997; Bungener et al.,
4 1999a, b; Franzaring et al., 2000), wild herbaceous plants (Bergmann et al., 1999; Danielsson
5 et al., 1999; Nussbaum et al., 2001), agricultural crops (Renaud et al., 1997; Ollerenshaw et al.,
6 1999; Benton et al., 2000; Heagle and Stefanski, 2000; Fumagalli et al., 2001; Elagoz et al.,
7 2002; Nali et al., 2002; Kollner and Krause, 2003), horticultural shrubs and trees (Hormaza
8 et al., 1996; Findley et al., 1997), and forest trees (Paakkonen et al., 1997; Volin et al., 1998; Pell
9 et al., 1999; Bortier et al., 2000; Postiglione et al., 2000; Landolt et al., 2000; Zhang et al., 2001;
10 Oksanen and Rousi, 2001; Matsumura, 2001; Guidi et al., 2001; Momen et al., 2002; Nali et al.,
11 2002; Saitanis and Karandinos, 2002). These studies have shown a wide range of responses to
12 O3, from growth stimulation by a few species such as Festuca ovina L. (Pleijel and Danielson,
13 1997) and Silene dioica and Chrysanthemum leucanthemum (Bungener et al., 1999) to
14 significant growth reduction, depending on environmental conditions and exposure dose.
15 While determining the explanation for differences in species sensitivity to O3 remains one
16 of the challenges facing plant biologists (Pell et al., 1999), a number of hypotheses have been
17 suggested. Reich (1987) proposed that variation in O3 sensitivity could be explained by variation
18 in total uptake of the gas. Others have suggested that (1) fast-growing species are more sensitive
19 than slower growing ones (Bortier et al., 2000), (2) overall O3 sensitivity may be closely linked
20 to root responses to O3 (Warwick and Taylor, 1995) or (3) the relative ability of species to
21 detoxify O3-generated reactive oxygen free radicals may determine O3 sensitivity (Alscher et al.,
22 1997; Pell et al., 1999). Volin et al., (1998) suggest that the relative rate of stomatal
23 conductance and the photosynthesis rate at a given conductance both contribute strongly to
24 determining species sensitivity to O3. Likely, there is more than one mechanism determining
25 sensitivity, even in a single species.
26 Within a given species, individual genotypes or populations can also vary significantly in
27 O3 sensitivity (Environmental Protection Agency, 1996). For example, the intraspecific
28 variation in O3 sensitivity for was a factor of two for Phleum prateme (Danielsson et al., 1999)
29 and Trifolium repens L. (Postiglione et al., 2000). A similar range of intraspecific variations in
30 O3 responses was demonstrated for clonal differences in Betulapendula by Paakkonen et al.
31 (1997) and Primus serotina (Lee et al., 2002). These examples of wide ranges within species
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1 responses suffice to show that caution should be taken when ranking species categorically as
2 having an absolute degree of O3 sensitivity (Davison and Barnes, 1998).
3
4 9.4.2.1 Genetic Basis of O3 Sensitivity
5 Plant response to ozone is determined by genes that are directly related to oxidant stress
6 and to an unknown number of genes that are not specifically related to oxidants. The latter
7 includes genes that control leaf and cell wall thickness, stomatal conductance and the internal
8 architecture of the air spaces. Although there is currently a great emphasis on individual
9 antioxidants that can be manipulated by molecular methods, the challenge is to determine the
10 relative contributions of all of the components to plant response and to understand the interplay
11 between them. Recent studies using molecular biological tools are beginning to increase the
12 understanding of O3 toxicity and differences in O3 sensitivity.
13 While much of the research in developing the understanding of O3 responses has been
14 correlative in nature, recent studies with transgenic plants have begun to positively verify the
15 role of various genes and gene products in O3 tolerance. Overexpressing MnSOD in chloroplasts
16 increased O3 tolerance in transgenic tobacco plants (Van Camp et al., 1994) provided the first
17 definitive proof of antioxidants key role in O3 tolerance. Subsequently, Broadbent et al. (1995)
18 showed that overexpression of pea glutathione reductase simultaneously in both chloroplasts and
19 mitochondria of transgenic tobacco enhanced O3 tolerance. Similarly, increased O3 tolerance to
20 O3-induced foliar necrosis was shown for transgenic tobacco plants Overexpressing the cytosolic
21 Cu/Zn-SOD gene (Pitcher and Zilinskas, 1996). Transgenic tobacco plants expressing antisense
22 RNA for cytosolic ascorbate peroxidase, which reduces ascorbate peroxidase production,
23 showed increased susceptibility to O3 injury suggesting a key role in O3 tolerance for the
24 antioxidant ascorbate peroxidase (Orvar and Ellis, 1997).
25 The consensus among molecular studies of O3 sensitivity is pointing to O3 triggering
26 salicylic acid, ethylene, and jasmonic acid and that the signaling of these molecules determines,
27 in some cases, the O3 susceptibility of plants (Overmyer et al., 2000; Rao and Davis, 1999; Rao
28 et al., 2000; Langerbartels et al., 2002; Moeder et al., 2002; Tamaoki et al., 2003; Vahala et al.,
29 2003a, b). Increased levels of jasmonic acid production in O3-tolerant compared to O3-sensitive
30 plants has been shown for Arabidopsis (Overmyer et al., 2000) and Populus (Koch et al., 1998,
31 2000). Blockage of ethylene production by using antisense methods with ACC synthase and
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1 ACC oxidase suggest strongly that ethylene synthesis and perception are required for H2O2
2 production and cell death following O3 exposure of Lycopersicon esculentum (Moeder et al.,
3 2002). Ethylene signaling may have multiple roles in O3 tolerance determination as was
4 demonstrated recently by Vahala et al., (2003a, b) who found that in Populus tremula x
5 P. tremuloides hybrid clones differing in O3 sensitivity, ethylene accelerated leaf senescence in
6 sensitive plants under low O3, but under acute O3, ethylene seemed to be required for protection
7 from cell death.
8 While changing expressions of single antioxidant genes has proven very useful in
9 identifying possible mechanisms of O3 sensitivity and tolerance (Kuzniak, 2002), it should be
10 noted that some studies of transgenic plants with enhanced antioxidant production have not
11 resulted in increased O3 tolerance (Saji et al., 1997; Torsethaugen et al., 1997; Strohm et al.,
12 1999, 2002). Clearly, ethylene production plays a role in O3 sensitivity but the role of various
13 antioxidants in O3 tolerance regulation are yet to be fully elucidated (Wellburn and Wellburn,
14 1996). It is unlikely that single genes are responsible for O3 tolerance responses, except in rare
15 exceptions (Engle and Gabelman, 1966). Regulation of stomatal opening and leaf structure
16 (Bennett et al., 1992; Elegoz and Manning, 2002) are also likely to play key roles in O3 tolerance
17 in plants. Newly developing opportunities to examine simultaneous regulation of larger numbers
18 of genes will likely yield more clarification of genes controlling O3 tolerance (Desikan et al.,
19 2001; Matsuyama et al., 2002).
20 Attempts to demonstrate conclusively changes in antioxidant and protective pigments for
21 O3 sensitive and tolerant mature trees growing in the field have largely been unsuccessful (Tausz
22 et al., 1999a, b). However, evidence for antioxidant expression differences contributing to
23 differences in O3 sensitivity of four-year-old Populus tremuloides trees has been found
24 (Wustman et al., 2001).
25
26 9.4.3 Environmental Biological Factors
27 The biological factors within the plant's environment that may influence its response to O3
28 directly or may be influenced to the advantage or disadvantage of plants encompass insects and
29 other animal pests, diseases, weeds and other competing plant species. Although such
30 interactions are ecological in nature, those involving individual pests, plant pathogens, or weeds,
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1 or agricultural crop or forest tree species are considered in this section. More complex
2 ecological inter-species interactions are dealt with in Section 9.5.
3 The different types of biological factors are dealt with separately, as in the 1996 O3 AQCD
4 (U.S. Environmental Protection Agency, 1996). Still, it is important to recognize certain general
5 features of relationships of plants with biological components of their environment:
6 • Successful infestation or infection involves complex interactions among the target or host
species, the causal organism and environmental factors.
7 • Infestations and infections may co-occur.
8 • The successful development and spread of a pest, pathogen or weed require favorable
environmental factors.
9 • Significant losses to crops and forest trees result from pests and pathogens.
10 • Significant losses to crops and seedling trees result from weed competition.
11 Ozone and other photochemical oxidants may influence the severity of a disease or
12 infestation by a pest or weed, either by direct effects on the causal species, or indirectly by
13 affecting the host, or both. In addition, the interaction between O3, a plant and a pest, pathogen
14 or weed may influence the response of the target host species to O3. A perceptive overview of
15 the possible interactions of O3-exposure with insect pests and fungal diseases has been provided
16 by Jones et al. (1994), based on a model system involving two insects and two pathogens
17 affecting cottonwood (Populus deltoides). Their study also included effects on the
18 decomposition of leaf litter.
19 In contrast to detrimental biological interactions, there are mutually beneficial relationships
20 or symbioses involving higher plants and bacteria or fungi. These include (1) the nitrogen-fixing
21 species Rhizobium and Frankia that nodulate the roots of legumes and alder, and (2) the
22 mycorrhizae that infect the roots of many crop and tree species, all of which may be affected by
23 exposure of the host plants to O3.
24 In addition to the interactions involving animal pests, O3 may also have indirect effects on
25 higher herbivorous animals, e.g., livestock, due to O3-induced changes in feed quality.
26
27 9.4.3.1 Oxidant-Plant-Insect Interactions
28 The 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996) stressed the variability
29 in the reported effects of O3 on host plant-insect interactions. Since relatively few plant-insect
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1 systems have been studied, few consistent patterns of response have emerged, as noted in other
2 reviews such as those of Colls and Unsworth (1992), Heliovaara and Vaisanen (1993), Whittaker
3 (1994), Docherty et al. (1997) and, most recently, Fluckiger et al. (2002).
4 None of the studies reported in the past decade have clarified the situation in terms of
5 clearly consistent effects. A 1997 review by Docherty et al. (1997), for example, examined
6 17 reports of studies of aphid species on a range of hosts and classed the O3 effects on aphid
7 performance as: 35% positive; 41% negative; and 24% showing no significant effect.
8 A tabulation of 19 studies by Fluckiger et al. (2002) gave the corresponding figures: 42%,
9 21%, and 3 7%.
10 Other recent studies with the aphids Schizolachnuspineti and Cinarapinea on Scots pine
11 (Pinus sylvestris) and Cinarapilicornis on Norway spruce (Picea abies) have also yielded
12 variable results, with them suggesting that O3 enhances aphid density on pine and aphid
13 performance on spruce (Holopainen et al., 1997; Kainulainen et al., 2000a). In an earlier study
14 with Schizolachnus pineti on Scots pine, Kainulainen et al. (1994) had observed no significant
15 effects of O3-treatment on aphid performance. However, more recent observations of long-term
16 effects on aphid populations on aspen (Populus tremuloides) exposed to O3 in a FACE system
17 revealed that O3 significantly increased aphid populations and decreased the populations of
18 predatory insects (Percy et al., 2002).
19 The observations of Brown et al. (1993) and Jackson (1995) led Whittaker (1994) and
20 Brown (1995) to suggest that aphid response was dependent on ambient temperature as well as
21 the dynamics of O3 exposure; growth tended to be stimulated with maximum temperatures below
22 -20 °C but was reduced at higher temperatures. The present situation with plant-aphid responses
23 therefore remains confused and, although numerous suggestions have been offered to explain
24 specific findings, they are difficult to assemble into a coherent picture.
25 Variability has also been found with the interactions involving chewing insects. For
26 example, Lindroth et al. (1993) reported a small negative O3 effect (8% reduction) on the growth
27 of gypsy moth larvae (Limantra dispaf) on hybrid poplar (Populus tristis x balsamiferd) but no
28 effect when on sugar maple (Acer saccharum). Ozone resulted in reduced growth rate of the
29 larvae of the bug Lygus rugulipennis on Scots pine, but enhanced the growth of larvae of the
30 sawfly Gilpiniapallida (Manninen et al., 2000). Costa et al. (2001) observed no significant O3
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1 effects on the growth and fecundity of the Colorado potato beetle (Leptinotarsa decemlincatd)
2 on potato in greenhouse and field experiments.
3 Fortin et al. (1997), in a two-year study of forest tent caterpillar (Malacosoma disstrid) on
4 sugar maple, observed that O3 exposure only increased the growth rate of female larvae in one
5 year; 4th- and 5th-instar larvae also showed a feeding preference for treated foliage in that year.
6 However, studies based on open-air exposures of aspen (Populus tremuloides) indicated
7 O3-enhanced growth of M disstria in terms of pupal weight (Percy et al., 2002) and larval
8 performance (Kopper and Lindroth, 2003). Jackson et al. (2000) observed inconsistency in
9 studies on the larva of the tobacco hornworm (Manduca sexto) on tobacco (Nicotiana tabacum).
10 In one year, feeding on O3-treated foliage resulted in significantly greater larval weight, whereas
11 in a second year the increase was not statistically significant although survival was increased.
12 Also, oviposition by hornworm moths was increased if ambient O3 levels were increased by 70%
13 but fell back to normal in ambient O3 levels (Jackson et al., 1999).
14 Studies of the two-spotted spider mite (Tetranychus urticae) on white clover (Trifolium
15 repens) and peanut (Arachis hypogeae) by Heagle et al. (1994) and Hummel et al. (1998)
16 showed that, on peanut and an O3-sensitive clover clone, O3-exposure stimulated mite
17 populations. The lack of significant effects on mites on the O3-resistant clover clone suggests
18 that the responses were host-mediated.
19 With chewing insects and mites, there therefore appears to be a clearer indication of the
20 likelihood that increased insect performance will result from O3-induced changes in the host
21 plant, but negative effects continue to be reported, indicating that the response is probably also
22 being determined in part by other environmental, genetic, or temporal variables.
23 Reported O3-induced enhancement of attack by bark beetles (Dendroctonus brevicomis) on
24 Ponderosa pine (Pinusponderosd) has been suggested by Dahlsten et al. (1997) to be due to
25 greater brood development on injured trees, possibly related to decreased numbers of predators
26 and parasitoids. This view gains some support from the observation that O3 adversely affected
27 the searching behavior of a parasitoid, Asobara tabida, for larvae ofDrosophila subobscura
28 which led to fewer parasitized hosts (Gate et al., 1995). Such observations reveal another level
29 of complexity in the O3-plant-insect interrelationship: O3 may reduce the effectiveness of the
30 natural control of insect pests. The phenomenon is probably related to effects on olfactory cues,
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1 since it was shown by Arndt (1995) that O3 can affect fly behavior by modifying the pheromones
2 that cause fly aggregation.
3 These reports focus on the direct or indirect effects on the insect or mite feeding on foliage
4 previously or currently exposed to O3. They provide little if any information on effects on the
5 host plant other than qualitative references to the injury caused by the O3-exposure. Enhanced
6 pest development will ultimately lead to increased adverse effects on the hosts in the long term,
7 but the only report of an O3-plant-insect interaction directly affecting the host plant in the short
8 term still appears to be that of Rosen and Runeckles (1976). They found that infestation by the
9 greenhouse whitefly (Trialeurodes vaporariorum) sensitized bean plants (Phaseolus vulgaris) to
10 injury by otherwise non-injurious low levels of O3, leading to premature senescence of the
11 leaves.
12 The overall picture regarding possible O3 effects on plant-insect relations, therefore,
13 continues to be far from clear. Only a few of the very large number of such interactions that
14 may affect crops, forest trees and other natural vegetation have been studied. The trend
15 suggested in the 1996 criteria document that O3 may enhance insect attack has received some
16 support from a few recent studies. However, the variability noted in most of the studies makes it
17 clear that we are still far from being able to predict the nature of any particular O3-plant-insect
18 interaction or its magnitude or severity.
19
20 9.4.3.2 Oxidant-Plant-Pathogen Interactions
21 Plant diseases are caused by pathogenic organisms, e.g., fungi, bacteria, mycoplasmas,
22 viruses, and nematodes. Ozone impacts on disease are briefly discussed in earlier reviews by
23 Ayres (1991) and Colls and Unsworth (1992) and, more recently, by Fluckiger et al. (2002).
24 Biotic interactions with forest trees have been reviewed by Chappelka and Samuelson (1998);
25 and Sandermann (1996) and Schraudner et al. (1996) have summarized molecular similarities
26 and interrelationships between necrotic O3 injury to leaves and pathogen attack. A few recent
27 publications have added to our fragmented knowledge of O3-plant-disease interactions and the
28 mechanisms involved, but there appear to have been no reports to date of studies involving
29 mycoplasmal diseases.
30 The 1996 criteria document (U.S. Environmental Protection Agency, 1996) noted the
31 concept put forward by Dowding, (1988) "that pathogens and pests which can benefit from
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1 damaged host cells and from disordered transport mechanisms are enhanced by pollution insult
2 to their hosts, whereas those pathogens and other symbionts which require a healthy mature host
3 for successful invasion are depressed by pollutant stress to their host." The pathogens of the first
4 type are mostly facultative, necrotrophic, fungal parasites, whereas the second type are largely
5 obligate biotrophic fungi, bacteria and viruses. Based on this distinction, the majority of the
6 cases cited in the 1996 document supported Dowding's view, as have several more recent studies
7 summarized in Table 9-11. However, there are also some contradictions.
8 Most investigations have focused on the incidence and development of disease on plants
9 previously or concurrently exposed to O3, rather than on the corollary effect of disease on the
10 response to O3. In all of the studies of facultative pathogens and the nematode studies, exposure
11 to O3 tended to result in increased disease severity through increased spore germination or
12 increased fungal growth and development, although in the case of grey mold (Botrytis cinered)
13 on kidney bean (Phaseolus vulgaris) this was only observed on an O3-sensitive cultivar
14 inoculated with conidia (Tonneijck, 1994). Using mycelial inoculation, O3 reduced disease
15 development, but no satisfactory explanation was offered to account for the difference in
16 response. With leaf spot, Marssonina tremulae, on hybrid poplar (Populus trichocarpa x
17 balsamifera), low level exposures to O3 also increased disease (in agreement with theory) but
18 higher levels (200 ppb, 8h per day for 15 days) reduced conidial germination (Beare et al.,
19 1999b).
20 The situation with obligate biotrophic pathogens is less consistent. The effects on powdery
21 mildew (Sphaerotheca fulginea) on both bottle gourd (Lagenaria siceraria) and cucumber
22 (Cucumis sativa) resembled the situation with the necrotrophic poplar leaf spot disease
23 (Marssonina tremulae), since low O3 exposures increased disease severity (in disagreement with
24 theory) although higher levels decreased it. The decreased infection in the pea-powdery mildew
25 (Erysiphe polygon!) situation agrees with theory, but the situations with leaf rust (Melampsora
26 sp.) on poplar (Populus trichocarpa x balsamifera) or aspen (Populus tremuloides) do not.
27 However, these reports are in contrast to earlier reports included in the 1996 criteria document
28 (U.S. Environmental Protection Agency, 1996) of observations with other species of Erysiphe
29 (Tiedemann et al., 1991) and Melampsora (Coleman et al., 1987). In contrast to the recent report
30 of a synergism with Melampsora on poplar, infections caused by the other biotrophs
31 (Sphaerotheca, Erysiphe, Uromyces) reduced the severity of injury caused by O3 (in agreement
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3
Table 9-11. Interactions Involving O3 and Plant Pathogens
o
o
1
O
H
6
o
0
H
O
o
0
o
H
W
Host Plant
Obligate Biotrophs
Bottle gourd (Lagenaria siceraria)
Cucumber (Cucumis saliva)
Pea (Pisum sativum)
Aspen (Populus tremuloides)
Hybrid poplar (Populus trichocarpa
x balsamifera)
Broad bean (Viciafaba)
Facultative Necrotrophs
Kidney bean (Phaseolus vulgaris)
Scots pine (Pinus sylvestris)
Pathogen
Powdery mildew
(Sphaerotheca fulginea)
Powdery mildew
(Sphaerotheca fulginea',
Powdery mildew
(Erysiphe polygoni)
Leaf rust (Melampsora
medusae f. sp.
tremuloidae)
Leaf rust (Melampsora
larici-populina or
M. allii-populina)
Bean rust (Uromyces
viciae-fabae)
Grey mold
(Botrytis cinerea)
Grey mold
(Botrytis cinerea)
White mold
(Sclerotinia sclerotiorum)
Annosus root and butt rot
(Heterobasidion annosum)
Effect of O3 on Disease
Increased in 50ppb O3;
decreased in 100+ppb
Increased in 50ppb O3;
decreased in 100+ppb
Decreased infection
Increased severity
Increased infection and
severity
Not reported
Increased from conidia on O3-
sensitive cultivar; decreased
from mycelium
Increased infection
Increased infection
Increased development*
Effect of Disease on O3 Response
Decreased; partial protection
Synergistic increase in 50ppb O3;
antagonistic decrease in 100+ppb;
partial protection
Decreased; partial protection
Not reported
Increased sensitivity (synergistic)
Decreased; partial protection
Not reported
Not reported
Not reported
Not reported
Reference
Khan and
Khan(1998a)
Khan and
Khan (1999)
Rusch and
Laurence
(1993)
Karnosky
et al. (2002)
Beare et al.
(1999a)
Lorenzini
et al. (1994)
Tonneijck
(1994)
Tonneijck and
Leone (1993)
Tonneijck and
Leone (1993)
Bonello et al.
(1993)
-------
3
Table 9-11 (cont'd). Interactions Involving O3 and Plant Pathogens
to
o
o
VO
oo
DRAFT-DO >
0
H
O
Host Plant
Facultative Necrotrophs (cont'd)
Loblolly pine (Pinus taeda)
Hybrid poplar (Populus deltoides x
nigra)
Hybrid poplar (Populus trichocarpa
x balsamifera)
Wheat (Triticum aestivum)
Nematodes
Tomato (Lycopersicon esculentum)
* Increase completely countered by
Pathogen
Pitch canker
(Fusarium subglutinans)
Canker (Septoria musiva
[=Mycosphaerella
populinum})
Leaf spot (Marssonina
tremulae)
Blotch (Septoria
nodorum)
Tan spot (Pyrenophora
tritici-repentis)
Root-knot nematode
(Meloidogyne incognita)
Effect of O3 on Disease Effect of Disease on O3 Response
Increased development Increased sensitivity
Increased incidence Not reported
Increased spore germination Not reported
and lesion growth after
lOOppb O3 (30 days);
decreased germination after
200ppb (15 days)
Increased infection Not reported
Increased infection of Not reported
disease-susceptible genotypes
Increased development Increased foliar injury; reduced plant
growth (synergistic)
Reference
Carey and
Kelley (1994)
Woodbury
et al. (1994)
Beare et al.
(1999b)
Tiedemann
and Firsching
(1993)
Sah et al.
(1993)
Khan and
Khan (1997,
1998a)
mycorrhizae (Hebeloma crustuliniforme).
O
HH
H
W
-------
1 with numerous earlier reports), but only at high O3 exposures in the case of Sphaerotheca on
2 cucumber. At low exposure levels, the disease and O3 acted synergistically. The only other
3 recent observation of such disease-related synergisms are the nematode-tomato reports of (Khan
4 and Khan, 1997, 1998b).
5 It is therefore clear that the type and magnitude of exposure to O3 plays an important role
6 in determining both the response of the disease organism and of the host.
7 No recent studies involving interactions between O3 and bacterial diseases appear to have
8 been reported. With regard to viruses, a laboratory study by Yalpani et al. (1994) added to
9 several reports of O3 decreasing the severity of tobacco mosaic virus infection of tobacco,
10 Nicotiana tabacum; and Jimenez et al. (2001) reported that previous exposure to O3 resulted in
11 increased adverse effects on tomato yield attributed to several virus diseases.
12 Similarities between the sensitivities of different cultivars or clones to O3 and to specific
13 diseases have been noted. For example, Sah et al. (1993) found that the severity of injury caused
14 by tar spot and standard O3 exposures of 12 wheat cultivars were closely correlated (R2 = 0.986).
15 Such similarities appear to have a mechanistic basis, since several studies have noted similarities
16 in the molecular and biochemical changes that occur in plants infected with pathogens and plants
17 exposed to O3. Schraudner et al. (1992), Ernst et al. (1992), Eckey-Kaltenbach et al. (1994a,b),
18 Yalpani et al. (1994) and Bahl et al. (1995) have presented evidence that exposures to O3 result
19 in responses such as increased levels of salicylic acid, the signaling agent for increased induced
20 resistance to pathogens. This, in turn, leads to activation of the genes that encode defense
21 proteins, including the so-called pathogenesis-related proteins. The induction of such proteins
22 might account for the decreased infection with Sphaerotheca and Melampsora at higher O3
23 exposures but does not account for increased infections seen at lower exposure levels. The issue
24 is discussed more fully by Sandermann (1996) and Schraudner et al. (1996). More recently
25 Sandermann has extended the theory relating O3 and disease by suggesting that, because of O3
26 "memory effects" in affected host plants that may persist over weeks or months, analysis for
27 various induced biomarkers of gene activation may provide a useful tool for improving our
28 ability to predict the outcome of O3-plant-pathogen interactions (Sandermann, Jr., 2000).
29 There have been no reports of O3 studies with mixed infections by pathogens, but the
30 complete suppression of Heterobasidion butt and root rot of Scots pine by the mycorrhizal
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1 symbiont Hebeloma crustuliniforme indicates the possibility of interactions involving more than
2 one fungus (see Section 9.4.4.3.3 below).
3 In summary, our understanding of oxidant-plant-disease interactions is far from complete.
4 However, a combined tabulation of the evidence presented in the 1996 O3 AQCD (U.S.
5 Environmental Protection Agency, 1996) and that noted in Table 9-11 leads to the following
6 summary of O3 effects on plant diseases and corollary effects of infection on plant response
7 to O3, as indicated by number of studies showing increases or decreases in disease or
8 susceptibility:
9 For obligate biotrophic fungi, bacteria, nematodes:
10 O3 increased disease: 9 Increased susceptibility to O3: 3.
11 O3 decreased disease: 15 Decreased susceptibility to O3: 9.
12 F or facultative necrotrophi c fungi:
13 O3 increased disease: 25 Increased susceptibility to O3: 2.
14 O3 decreased disease: 3 Decreased susceptibility to O3: 4.
15 Thus, although O3 may reduce the severity but not the incidence of some of the diseases caused
16 by the obligate pathogens, the evidence overall indicates that with most diseases, their severity is
17 more likely to be increased by O3 than not. However, the actual consequences will be specific to
18 the disease and level of exposure, and, most importantly, will be determined by environmental
19 suitability and epidemiological requirements for disease to develop. Conversely, some evidence
20 suggests that infection by obligate pathogens may confer some degree of "protection" against O3,
21 a dubious benefit from the plant's point of view.
22
23 9.4.3.3 Oxidant-Plant-Symbiont Interactions
24 No further studies have appeared regarding O3 effects on the important bacterial symbiont
25 of legumes, Rhizobium, since those summarized in the 1996 document (U.S. Environmental
26 Protection Agency, 1996). Hence, our present understanding is that, although relatively high
27 levels of exposure (> 200 ppb) can result in severe (> 40%) reductions in nodulation (and
28 therefore nitrogen-fixation) on soybean roots, lesser reductions in nitrogen-fixing may be caused
29 by lower exposures. However, the data are inadequate to attempt to define any quantitative
30 exposure-response relationships.
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1 There have been a few recent reports on O3-plant-mycorrhizae interrelationships. These
2 have mostly involved seedlings of coniferous tree species. A transient O3-induced stimulation of
3 mycorrhiza on Scots pine roots reported by Kasurinen et al. (1999) was not observed in a later
4 study by Kainulainen et al. (2000b). Studies of the mycorrhiza Paxillus involutus on birch
5 (Betulapenduld) seedlings showed that, although O3 reduced mycorrhizal growth rate, it led to
6 greater extension growth which in turn resulted in greater mycorrhizal infection of neighboring
7 Aleppo pine (Pinus halepensis) seedlings (Kytoviita et al., 1999). However, O3 reduced nitrogen
8 acquisition by P. halepensis from its mycorrhizal symbiont (Kytoviita et al., 2001). The
9 complex interrelationships that may occur in the rhizosphere were revealed by the observation
10 (Bonello et al., 1993) that the mycorrhiza Hebeloma crustuliniforme could overcome the
11 O3-stimulated severity of root rot on Scots pine (Pinus sylvestris) caused by the fungus
12 Heterobasidion annosum (noted in Section 9.4.4.3.2).
13 In summary, the available evidence is far too fragmented and contradictory to permit
14 drawing any general conclusions about mycorrhizal impacts. The negative effects of O3 on
15 mycorrhizae and their functioning that have been reported have not necessarily been found to
16 lead to deleterious effects on the growth of host plants. Thus, little has changed from 1991 when
17 Dighton and Jansen asked: " Atmospheric Pollutants and Ectomycorrhizae: More Questions
18 than Answers?" (Dighton and Jansen, 1991). Because of their important roles in ecosystems,
19 mycorrhizae are further discussed in Section 9.5 below.
20
21 9.4.3.4 Oxidant-Plant-Plant Interactions: Competition
22 Plant competition involves the ability of individual plants to acquire the environmental
23 resources needed for growth and development: light, water, nutrients and space. Intraspecific
24 competition involves individuals of the same species, typically in monocultural crop situations,
25 while interspecific competition refers to the interference exerted by individuals of different
26 species on each other when they are in a mixed culture.
27 In cropping situations, optimal cultural practices for row spacing and plant density/row
28 tend to balance the negative effects of intraspecific competition and the goal of maximum yield.
29 Although interspecific competition is agriculturally undesirable when it involves weak
30 infestations, the use of mixed plantings may be agriculturally deliberate, e.g., grass-legume
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1 mixtures used for pasture or forage. In natural plant communities, monocultures are rare, and
2 complex interspecific competition is the norm.
3 Although weak competition is the largest global cause of crop losses, little is known about
4 the impact of O3 on crop-weed interactions. The topic does not appear to have been investigated
5 in recent years. We can only speculate as to the possible consequences of O3 exposure on weed
6 competition based on our limited understanding of the effects on a few, mostly two-component
7 mixtures of cultivated species.
8 The tendency for O3-exposure to shift the biomass of grass-legume mixtures in favor of
9 grass species, reported in the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996) has
10 been confirmed by recent studies. In a ryegrass (Lolium perenne) + clover (Trifolium repens)
11 mixture grown in an open-air fumigation system, clover growth was impaired by extended
12 exposures to above-ambient O3, leaving patches for weed invasion (Wilbourn et al., 1995).
13 An open-top chamber study by Nussbaum et al. (1995b) using the same species confirmed the
14 greater effect on clover but observed that the magnitude of the effect depended highly on the
15 pattern of O3-exposures over extended growing periods. Low-level exposures shifted species
16 composition in favor of Lolium., but exposures to higher peak O3 levels depressed total mixture
17 yield. With an alfalfa (Medicago saliva) + timothy (Phleumprateme) mixture, Johnson et al.
18 (1996a) noted that O3 caused decreased alfalfa root growth and increased timothy shoot growth
19 and height. Nussbaum et al. (2000a) reported that, with increased exposure to O3, well-watered
20 red clover (Trifolium prateme) plants suffered from increased competition from the grass,
21 Trisetumflavescens, but the O3 exposure also negatively effected grass growth, depressing
22 overall total yield. However, a greater adverse effect on Trisetum resulted from O3-induced
23 increased competition when grown with brown knapweed (Centaureajacea)., a weed species.
24 Andersen et al. (2001) demonstrated the potential for competition and O3-exposure to work
25 together to affect the growth of tree seedlings. Ozone had no direct adverse effect on pine
26 growth in a 3-year study of ponderosa pine (Pinusponderosa} seedlings grown in mesocosms
27 with three densities of blue wild-rye grass (Elymus glaucus), but the O3-increased competitive
28 pressure of the grass caused a major reduction in pine growth.
29 Three studies have been reported on more complex plant associations. Ashmore and
30 Ainsworth (1995) studied mixed plantings of two grasses, Agrostis capilaris and Festuca rubra,
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1 with two forbs16, Trifolium repens (a legume) and Veronica chamaedrys, exposed to O3 in
2 open-top chambers (OTCs). The proportion of forbs, Trifolium in particular, declined, especially
3 when cut at biweekly intervals. In a related study, Ashmore et al. (1995) used artificial mixtures
4 of grasses and forbs and transplanted swards of native calcareous grassland species and found
5 that, regardless of whether total biomass was adversely affected by exposures to O3, higher
6 exposures progressively shifted species composition, usually at the expense of the forb species.
7 The observed shifts in competitive balance in favor of grasses is consistent with observations
8 that many grass species are less sensitive to O3 than forbs. However, as previously shown by
9 Evans and Ashmore (1992), knowledge of the relative sensitivities to O3 of the component
10 species grown in isolation or in monoculture does not always predict the impact of O3 on the
11 components in a mixed culture.
12 Barbo et al. (1998) exposed an early successional plant community to O3 in OTCs for two
13 growing seasons. Ozone decreased community structure features such as height of canopy,
14 vertical canopy density (layers of foliage), and species diversity and evenness. Surprisingly,
15 blackberry (Rubus cuneifolius), a species considered to be O3-sensitive, replaced sumac (Rhus
16 copallind) canopy dominance. Barbo et al. (2002) also demonstrated the role of competition in
17 determining the impact of O3 on loblolly pine (Pinus taedd). They reported that the increased
18 growth of natural competitors in OTCs using charcoal-treated air to reduce the ambient O3
19 concentrations resulted in decreased pine growth. They noted that this is contrary to the
20 frequently reported increased growth observed in reduced O3 levels in the absence of
21 interspecific competition.
22 Competitively disadvantaged trees of four clones of aspen (Populus tremuloides) exposed
23 to 1.5* ambient O3 in a FACE facility over 4 years were proportionately more adversely affected
24 by O3 than competitively advantaged or neutral trees (McDonald et al., 2002). However, one
25 clone of the disadvantaged trees demonstrated enhanced growth.
26 In summary, our present knowledge of how O3 may affect the competitive interspecific
27 plant-plant relationships typifying the agricultural and natural worlds is very limited. However,
28 as noted in the 1996 AQCD (U.S. Environmental Protection Agency, 1996), "the development
29 and use of field exposure systems have permitted many recent studies of crop species to be
30 conducted at normal planting densities and hence have incorporated intraspecific competition as
16Forb: any non-grassy herbaceous species on which animals feed.
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1 an environmental factor." Such facilities were used in most of the studies of interspecific
2 competition discussed above. But we are still a long way from being able to bridge the gap
3 between small model competing systems and the realities of natural ecosystem complexity.
4
5 9.4.4 Physical Factors
6 The physical features of a plant's aerial and edaphic environments exercise numerous
7 controls over its growth and development. Thus, many of their effects may be modified by
8 exposure to atmospheric oxidants and, alternatively, plants may modify responses to such
9 exposures. As in the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996), this
10 section focuses on the defining features of plant microclimate: light, temperature, relative
11 humidity (HR, or saturation vapor pressure deficit), and the presence and availability of water,
12 especially in the soil. Monteith and Elston (1993) suggested that light energy and mass of water
13 should be viewed as climatic resources, and the other two elements (temperature and saturation
14 vapor pressure deficit) as rate modifiers that determine how fast the resources are used. The
15 modifications of plant response by physical environmental factors has recently been reviewed by
16 Mills (2002).
17 Another physical feature of the microclimate, wind and air turbulence, which affects the
18 thicknesses of the boundary layers over leaves and canopies and, hence, affects gas exchange
19 rates (including the fluxes of O3 and other oxidants into the leaves) is discussed elsewhere
20 (Section 9.3).
21 Physical features of the environment are also important components of larger-scale
22 regional and global climates. However, the following discussions are confined to issues related
23 to individual factors at the plant level; meso-scale effects are reviewed in Section 9.4.4.8, which
24 addresses the issues of climate change interactions.
25
26 9.4.4.1 Light
27 Plants are the primary producers of biomass on the planet through their ability to capture
28 light energy (by the process of photosynthesis) and convert it to the many forms of chemical
29 energy that sustain their own growth and that of secondary consumers and decomposers. Light
30 intensity is critical since the availability of light energy (a resource, sensu [Monteith and Elston,
31 1993]) governs the rate at which photosynthesis can occur, while light duration (i.e.,
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1 photoperiod) profoundly effects development in many species. Although light quality (i.e., the
2 distribution of incident wavelengths) may also affect some physiological plants processes, there
3 is no evidence to indicate that such effects are of relevance to concerns over oxidant pollution,
4 except at the short wavelengths of UV-B. This topic is discussed in the context of climate
5 change in Section 9.4.8, and as a stress factor per se affected by atmospheric O3 in Chapter 8.
6 However, as noted above and in the 1996 O3 AQCD (U.S. Environmental Protection Agency,
7 1996), none of the features is controllable in natural field situations. A brief discussion of light
8 intensity-O3 interactions is included in the review by Chappelka and Samuelson (1998).
9 The conclusion in the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996) that
10 low light intensities and short photoperiods tended to increase susceptibility to foliar O3-injury
11 may still be valid, but this may or may not translate into adverse effects on growth. For
12 example, Tjoelker et al. (1993) found that, when seedlings of sugar maple (Acer sacchamm),
13 a shade-tolerant species, were grown in 7% full sunlight, O3 reduced shoot and root growth, but
14 had no significant effect in 45% sunlight (a 6-fold increase). In contrast, the reverse was
15 observed with a shade-intolerant hybrid poplar (Populus tristis x balsamifera\ with the greater
16 impact of O3 occurring in the higher light intensity treatment.
17 The greater sensitivity of maple in low light has also been confirmed in other studies.
18 Tjoelker et al. (1995) noted a greater O3-induced inhibition of photosynthetic CO2 assimilation in
19 shaded leaves than in leaves in full sunlight. However, in the absence of differences in stomatal
20 conductance, the effect was considered to be independent of O3 flux; it appeared to be a
21 consequence of reduced chlorophyll contents and quantum efficiencies induced by O3.
22 In contrast, Back et al. (1999), who also observed a greater inhibition of net photosynthesis by
23 O3 in shaded leaves, reported decreased stomatal conductance. Although reduced conductance
24 might suggest reduced O3 flux and, therefore, decreased adverse effects, the authors concluded
25 that the effects of reduced conductance was offset by long-term changes in leaf structure, leading
26 to less densely packed mesophyll cells and greater internal air space within the leaves.
27 Morphological differences between lower and upper crown leaves of black cherry (Prunus
28 serotina) have been suggested as the basis for the greater O3-susceptibility of the lower crown
29 leaves (Fredericksen et al., 1995). Back et al. (1999) also observed accelerated foliar senescence
30 induced by O3 on shaded leaves, a response also noted by Topa et al. (2001). Sensitivity to O3
31 was found to be increased in shade- but not sun-leaves of shade-tolerant red oak (Quercus rubrd)
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1 (Samuelson and Edwards, 1993). Similarly, Mortensen (1999) observed that seedlings of
2 mountain birch (Betulapubescens) grown in 50% shade suffered greater foliar injury from O3
3 than those grown in full sunlight.
4 Not all shade-intolerant species exhibit greater reductions in photosynthesis and growth
5 due to O3 when grown in full sunlight. Higher than ambient levels of O3 failed to inhibit
6 photosynthesis in leaves of shade-intolerant yellow poplar (Liriodendron tulipifera) grown in
7 nearly full sunlight (Tjoelker and Luxmoore, 1991). Greater foliar injury in the lower, shaded
8 leaves of shade-intolerant black cherry (Primus serotind) trees and saplings, was attributed
9 (Fredericksen et al., 1996a) to higher stomatal conductance and greater O3 uptake relative to net
10 photosynthetic rate. However, in a 3-year study of Norway spruce seedlings (Picea abies) in
11 OTCs. Wallin et al. (1992) observed that photosynthetic efficiency was more adversely affected
12 by O3 in high than in low light.
13 The suggestion of greater sensitivity to O3 of shade-tolerant species in low-light conditions
14 and the greater sensitivity of shade-intolerant species in high light is somewhat of an over-
15 simplification when dealing with mature trees, for which light intensity varies considerably
16 within the canopy because of shading. Chappelka and Samuelson (1998) noted that the
17 interaction between sensitivity to O3 and the light environment in forest trees is further
18 complicated by developmental stage, with seedlings, saplings, and mature trees frequently giving
19 different results. Topa et al. (2001) also cautioned that O3 effects on leaf-level photosynthesis
20 may be poor predictors of the growth responses of sugar maple in different light environments.
21 In high-light intensities, many species exhibit some degree of photoinhibition of the
22 photosynthetic process through the overloading of the mechanisms that protect the
23 photosynthetic reaction centers in the chloroplasts. Guidi et al. (2000) reported complex
24 interactions between high-light intensities (inducing photoinhibition) and O3 exposures in kidney
25 bean (Phaseolus vulgaris) with high intensities tending to enhance the detrimental effect of O3
26 on photosynthesis. One of the studies in the extensive European Stress Physiology and Climate
27 Experiment-wheat (ESPACE-wheat) program (Bender et al., 1999), conducted in 1994 and
28 1995, included an investigation of the effects of climatic variables on yield response to O3 using
29 two simulation models, AFRCWHEAT2-O3 and LINTULCC (Ewart et al., 1999; Van Oijen and
30 Ewart, 1999). Among the observed trends, it was noted that relative yield loss of wheat due to
31 elevated O3 tended to increase with light intensity. In contrast, Balls et al. (1996) used ANNS to
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1 investigate microclimatic influences on injury caused by O3 to clover (Trifolium subterraneimi)
2 and found that, especially at mid-range cumulative O3 exposures (350 to 500 ppb-h), injury
3 tended to decrease with increasing light intensity. Similar observations by Davison et al. (2003)
4 of foliar injury to wild populations of cutleaf cone flower (Rudbeckia laciniatd) exhibiting a
5 range of PAR levels within their canopies led the authors to conclude that the variation in injury
6 symptoms observed was "unlikely to be due to differences in ozone flux and more likely to be
7 due to variation in light." Antonielli et al. (1997) found evidence indicating that the high
8 sensitivity of the bioindicator tobacco cultivar Nicotiana tabacum cv Bel-W3 is partly
9 determined by its high photosynthetic electron transport rates at high-light intensities, which
10 exceed the capabilities of the plant to dissipate energy and oxyradicals.
11 The 1996 O3 AQCD referred to the important role of light in controlling stomatal opening
12 and suggested that light duration (i.e., photoperiod) might dictate the actual uptake of O3 to some
13 degree. However, it should also be noted that Sild et al. (1999) found that clover plants could
14 suffer foliar injury even if they were exposed to O3 during the dark period of the day-night cycle,
15 when stomatal conductance is at its lowest.
16 A possible indirect effect of light intensity was noted by Reiling and Davison (1992) in
17 their study of the O3-tolerance ofPlantago major plants grown from seeds collected from
18 populations at 28 different sites in Britain. Ozone-tolerance, defined in terms of plant growth,
19 was found to be a function of both previous O3-exposure history and hours of bright sunshine
20 during the year before the seeds were collected. However, the authors cautioned that, since
21 tropospheric O3-formation is itself dependent upon irradiation, the observation does not
22 necessarily imply a direct effect of light intensity on the plants' response to O3.
23 The only recent studies concerning interactions with light quality appear to be those
24 involving O3 and UV-B as a component of climate change. These are dealt with in Section
25 9.4.4.8.2. The effects of photoperiod on response to O3 or the converse do not appear to have
26 received any recent attention.
27 Although the intensity, quality, and duration of light are not controllable in the natural
28 world, the interactions of O3 with light intensity, in particular, clearly have relevance to the
29 growth of shade-tolerant and shade-intolerant species in mixed forest stands. It appears that the
30 nature of light intensity-O3 interactions may depend upon the type of light environment to which
31 the species are best adapted, with increased light intensity increasing the sensitivity of light-
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1 tolerant and decreasing the sensitivity of shade-tolerant species to O3. Although there is
2 certainly some evidence to the contrary, this hypothesis is a reasonable summation of current
3 understanding with regard to O3-light intensity interactions.
4
5 9.4.4.2 Temperature
6 "Temperature determines the start and finish and rate and duration of organ growth and
7 development" (Lawlor, 1998). Such processes depend on fundamental physiological activities
8 that are mostly enzyme-mediated and whose kinetics are directly affected by temperature. Since
9 the processes of enzyme deactivation and protein denaturation also increase as temperatures rise,
10 each enzymatic process has a unique optimum temperature range for maximal function.
11 However, the optima for different processes within the plant vary appreciably and, hence, the
12 optimum temperature range for overall plant growth is one within which all of the individual
13 reactions and vital processes are collectively functioning optimally, not necessarily maximally.
14 Furthermore, individual features of plant development (e.g., shoot and root growth, flowering,
15 pollen tube growth, fruit set, seed development) have different specific optima, so that
16 differential responses to temperature occur, leading to temperature-induced developmental
17 changes. For example, despite increased assimilation, increased temperatures may result in
18 decreased grain yields of crops such as wheat, because the growing season is effectively
19 shortened by a more rapid onset of senescence (Van Oijen and Ewart, 1999).
20 Rowland-Bamford (2000) noted that a plant's response to temperature changes will depend
21 upon whether it is growing at its near optimum temperature for growth or its near maximum
22 temperature, and whether any increase in mean temperature results in temperatures rising above
23 the threshold for beneficial responses. Impairment by O3 of any process may be thought of as
24 being analogous to a downward shift below and away from the temperature optimum or an
25 upward shift above and away from the optimum. Since a temperature rise toward the optimum
26 would result in a rate increase, the combined effects of O3 and such an increase might neutralize
27 each other, while the effects of O3 and a decrease in temperature would likely be additively
28 negative. Above the optimum temperature, the situations would be reversed with the effects of
29 increased temperatures and O3 being additively negative, and decreasing temperatures
30 counteracting any negative effect of O3. Thus, it is difficult to generalize about the interactions
31 of temperature and O3 on overall plant responses such as growth in which the different
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1 temperature-rate relationships of different growth components are merged, because they depend
2 upon the relationship of any temperature changes to the optimum for a species.
3 Studies of the effects of temperature on the impact of O3 have increased recently because
4 of an increased need to understand the consequences of global warming as a component of
5 climate change. Direct interactions of temperature with O3 are reviewed here, but the issues are
6 addressed again in Section 9.4.8.1 in relation to changes in atmospheric CO2 levels.
7 The 1996 criteria document (U.S. Environmental Protection Agency, 1996) stressed the
8 interdependence of the temperature within the tissues of the leaf (where the various temperature-
9 sensitive processes occur) on three distinct components: the ambient air temperature, the heating
10 effect of incident infrared radiation during the photoperiod, and the evaporative cooling effect
11 caused by transpirational loss of water. It also cautioned that, especially in experiments using
12 controlled environment chambers, the effects of temperature could well be confounded with
13 those of humidity /vapor pressure deficit (VPD). Temperature and VPD are strongly interrelated,
14 and VPD plays an important role in regulating stomatal transpiration. Because of the role that
15 evaporative cooling plays in determining internal leaf temperatures, any factor that causes
16 stomatal closure and reduced conductance inevitably leads to increased leaf temperatures. Such
17 interactions add to the difficulties in distinguishing the effects of temperature from those of other
18 factors, as actual leaf temperatures are rarely measured and reported.
19 Despite these caveats, there is some evidence that temperature per se influences plant
20 response to O3. For example, in rapid-cycling Brassica (Brassica rapd) and radish (Raphanus
21 sativus), marked O3-inhibitions of growth were observed at low root temperatures (13 °C) but
22 not at 18°C (Kleier et al., 1998, 2001). With regard to air temperature, this was included in the
23 range of micrometeorological variables studied in several recent extensive field studies and was
24 found to have a significant effect on response to O3 in most cases. Ball et al. (1996) used ANNs
25 in an analysis of the growth of clover (Trifolium subterraneuni) and concluded that light and
26 VPD had greater influences than temperature on the visible injury response to O3. However, in
27 three studies with different cultivars of white clover (Trifolium repens), temperature was found
28 to be important to the growth response. Ball et al. (1998) exposed T. repens cv. Menna to
29 ambient O3 in OTCs at 12 European sites at a range of latitudes and altitudes from 1994 to 1996.
30 The impact of O3 on growth was determined as the ratio of growth with and without treatment
31 with the O3-protectant, EDU (see Section 9.2). Artificial neural network analysis showed that O3
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1 exposure (measured as the AOT40 index, see Section 9.4.6), VPD, and temperature were
2 consistently the three most important variables governing response to O3 over a range of
3 different ANN models. However, the authors did not describe the form of the O3-response
4 relationship with temperature. Similar observations were reported by Ball et al. (2000) and Mills
5 et al. (2000) for O3-sensitive and O3-tolerant clones of T. repens cv. Regal, grown at 14 to
6 18 European locations from 1995 to 1998. In both studies, the impact of O3 was measured as the
7 sensitive/tolerant growth (biomass) ratio. Although Ball et al. (2000) found temperature to be
8 less important than O3 exposure and VPD, Mills et al. (2000) found temperature to be the most
9 important input variable after O3 exposure (AOT40). In both cases, the adverse effect of O3
10 increased with increasing temperature.
11 A study of black cherry (Prunus serotind) seedlings and mature trees in Pennsylvania,
12 using micrometeorological variables aimed to predict O3 uptake, found temperature to be
13 unimportant (Fredericksen et al., 1996b), but in the study of populations ofPlantago major
14 referred to in Section 9.4.4.1, Reiling and Davison (1992) noted a weak, positive correlation
15 between mean temperature at the collection site and O3 tolerance (based on growth rate) of the
16 different populations. In contrast, Danielsson et al. (1999) collected genotypes ofPhleum
17 arvense from a wide range of Nordic locations and found a positive effect of temperature on the
18 growth of genotypes from locations wth higher summer temperatures, but sensitivity to O3 did
19 not vary systematically with geographic location.
20 Van Oijen and Ewart (1999) studied the effects of climatic variables on the response to O3
21 in the ESP ACE-wheat program, based on two distinctive simulation models (AFRCWHEAT2-
22 O3 and LINTULCC, [Ewart et al., 1999]) and noted that although the relative yield loss of wheat
23 due to elevated O3 tended to increase with temperature, the effect was of minor significance.
24 In contrast to the variable results obtained in studies of the effects of temperature on
25 response to O3, the corollary effect of O3 exposure on subsequent sensitivity to low temperature
26 stress, noted in the 1996 criteria document, is well recognized. In reviewing low temperature-O3
27 interactions, Colls and Unsworth (1992) noted that winter conditions produce three kinds of
28 stress: desiccation, chilling or freezing temperatures, and photooxidation of pigments. Of these,
29 they suggested that while the first two were important, the last may play a particularly significant
30 role because the "combination of high irradiance and low temperatures permits a buildup of free
31 radicals in leaf tissue, and these free radicals then attack chlorophyll." Chappelka and Freer-
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1 Smith (1995) suggested that the injury and losses to trees caused by this delayed impact of O3
2 may be equally or more important than the direct impacts of O3 on foliage of visible injury and
3 necrosis, or the disruption of key physiological processes such as photosynthesis. In this
4 context, the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996) referred to the
5 conceptual framework (Eamus and Murray, 1991), which is still valid: brief periods of mild
6 temperatures in the severest winters result in dehardening; O3 decreases frost hardiness per se,
7 but also increases the predisposition to dehardening; dehardening places O3-exposed trees at
8 greater risk from subsequent low temperatures. However, no quantified models of these effects
9 have yet appeared.
10 The 1996 criteria document also noted that O3 adversely affects cold hardiness of
11 herbaceous species. More recently, Foot et al. (1996, 1997) observed winter injury and
12 decreased growth in low-growing perennial heather Calluna vulgaris exposed to O3 (70 ppb,
13 8 h/day, 5 days/week for 6 months) during the winter (6.8 °C mean), but found no significant
14 effects of the same exposures during the summer (12.3 °C mean). Although Potter et al. (1996)
15 observed a similar situation with the moss Polytrichum commune, the reverse was found with the
16 moss Sphagnum recurvum.
17 In summary, unequivocal evidence exists that O3 causes sensitization to the adverse effect
18 of low temperatures, but there is no clear pattern to evidence regarding the effects of temperature
19 on O3 response. The many contradictory responses to temperature and O3 probably reflect our
20 lack of detailed knowledge of the temperature optima for the different growth components of the
21 studied species. The topic of temperature-oxidant interactions is revisited later in Section 9.4.4.8
22 in the context of global warming as a feature of climate change.
23
24 9.4.4.3 Humidity and Surface Wetness
25 The moisture content of the ambient air (or its VPD) is a rate modifier (sensu Monteith and
26 Elston, 1993) and an environmental regulator of stomatal conductance. Both of the previous
27 criteria documents (U.S. Environmental Protection Agency, 1986, 1996) concluded that the
28 weight of evidence indicated that high RH (= low VPD) tended to increase the adverse effects of
29 O3, principally because the stomatal closure induced in most situations by O3 is inhibited by high
30 RH, leading to increased O3 flux into the leaves.
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1 Recent reports have confirmed this role of RH. The studies by Ball et al. (1995, 1996,
2 1998) showed that VPD was an important determinant of O3-induced injury and reduced growth
3 in two species of clover, Trifolium repens cv. Menna and T. subterraneum. However, Mills et al.
4 (2000) found it to be unimportant in the case of T. repens cv. Regal. Such difference between
5 cultivars is not unexpected since considerable differences also occur among species and genera.
6 For example, Bungener et al. (1999a) studied 26 Swiss grassland species and found clear
7 evidence that O3 injury increased with decreased VPD (i.e., increased RH) in only eight species.
8 However, the 1995 data from the European cooperative study of O3 injury, which involved
9 28 sites in 15 countries and six crop species, led to the development of two 5-day critical-level
10 scenarios involving O3-exposure (calculated as the AOT40 index) and mean VPD (0930-1630h):
11 200 ppb-h at > 1.5 kPa, and 500 ppb-h at < 0.6 kPa (Benton et al., 2000).
12 With forest tree species, Fredericksen et al. (1996b) found significant correlations between
13 stomatal conductance of black cherry (Prunus serotina) leaves and RH (+ve) and VPD (-ve),
14 and studies on free-standing Norway spruce (Picea abies) and larch (Larix decidud) showed that
15 although ambient VPD was highly positively correlated with ambient O3 concentration,
16 increased VPD caused stomatal closure, reducing O3 uptake and impact (Wieser and Havranek,
17 1993, 1995).
18 Surface wetness may affect the response to O3 through its direct effects on deposition to
19 the surface and through changes in RH. Effects on the deposition of O3 have been reviewed
20 (Cape, 1996). A surface film of water on leaves was found to increase O3 deposition in four
21 studies involving field-grown grape (Vitis vinifera) (Grantz et al., 1995), red maple {Acer
22 rubruni) (Fuentes and Gillespie, 1992), deciduous forest dominated by largetooth aspen (Populus
23 grandidentata) and red maple (Fuentes et al., 1992), and clover-grass mixed pasture (Trifolium
24 pratense, Phleumpmtense, and Festucapratensis) (Pleijel et al., 1995). In each case, the
25 increased deposition could be attributed partly to an increased stomatal conductance through the
26 abaxial (lower) surface and partly to uptake into the aqueous film on the adaxial (upper) surface.
27 In contrast, decreased deposition was noted by Grantz et al. (1997) with field-grown cotton
28 (Gossypium hirsutum). Since cotton is amphistomatous, with functional stomata on both leaf
29 surfaces, it was suggested that, in this case, the water layer effectively sealed the adaxial surface
30 stomata, more than offseting any increase in conductivity of the stomata in the abaxial surface.
31 However, none of the studies investigated the consequences of the differences in deposition.
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1 Although it could be inferred that, with part of any increased deposition being the result of
2 increased O3 flux into the leaves, there would be the likelihood of increased O3 adverse effects,
3 as suggested by earlier studies (Elkiey and Ormrod, 1981) that, by misting bluegrass (Poa
4 pratemis) during exposure to O3, injury was significantly increased.
5 To conclude, the effects of high RH (low VPD) and surface wetness have much in
6 common, as they both tend to enhance the uptake of O3, largely through effects on stomata,
7 leading to increased impact.
8
9 9.4.4.4 Drought and Salinity
10 The 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996) concluded that the
11 available evidence clearly indicated that exposure to drought conditions could reduce the adverse
12 effects of O3 on the growth of herbaceous and woody plants, but it also noted that no quantitative
13 models of the O3-soil moisture deficit (SMD) interaction had yet appeared in print.
14 Nevertheless, the "protective" effect was inconsistent, and only appeared when SMD was
15 accompanied by high evaporative demand. Since that time, further studies have confirmed the
16 interaction, and simulation models have begun to appear. Mills (2002) has recently provided a
17 brief review of the topic.
18 With regard to herbaceous species, Vozzo et al. (1995) observed less O3-induced injury and
19 suppression of net photosynthesis and growth in water-deficient soybean (Glycine max) than in
20 well-watered plants. In several studies with wheat (Triticum aestivum), on the other hand,
21 although adverse effects of both O3 and SMD were noted, they were consistently additive
22 (Bender et al., 1999; Fangmeier et al., 1994b, 1994a; Ommen et al., 1999).
23 In attempting to model the stomatal conductance of wheat in relation to O3 and soil
24 moisture, Griiters et al. (1995) found that although O3-induced stomatal closure was enhanced by
25 SMD, reducing O3 uptake, the R2 of the overall model was only 0.40, indicating that other
26 significant factors or relationships were involved.
27 With regard to native vegetation, Bungener et al. (1999a) used mixed plantings of 24 Swiss
28 grasses, herbs, and legumes and observed that, although O3-drought interactions were species-
29 specific, they tended to reflect stomatal functioning. They found that SMD reduced O3 injury in
30 two clovers (Trifolium repens and T. pratensis) and two grasses (Trisetum flavescens and
31 Bromus erectus\ but noted no interactions in the other 20 species. With relative growth rate as
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1 the measure of response to O3, interactions with SMD were noted in only three species:
2 Trifolium repens and two weedy herbs, Knautia arvense and Plantago lanceolata (Bungener
3 et al., 1999b). Although this variability in response among species was noted in the review by
4 Davison and Barnes (1998), they also pointed out that in severely draughted regions of Europe,
5 notably in Greece and Spain, O3-induced injury and growth reductions were common on many
6 (usually irrigated) crops, but there were virtually no records of injury symptoms in wild species.
7 Thus, the situation with herbaceous species is essentially unchanged from 1988 when
8 Heagle et al. summarized the extensive NCLAN experiments that incorporated water stress as a
9 variable: "SMD can reduce the response of crops to O3 under some conditions but not under
10 other conditions. Probably the occurrence of O3 by SMD interactions was dependent on the
11 degree of SMD-induced plant moisture stress."
12 With regard to trees, O3 interactions with soil water availability have been discussed in
13 several recent reviews: Chappelka and Freer-Smith (1995), who focused on O3-induced
14 predisposition to drought stress; Johnson et al. (1996b); Chappelka and Samuelson (1998); and
15 Skarbyetal. (1998).
16 Several recent studies with conifers have yielded mixed results. No interactions with
17 drought were observed by Broadmeadow and Jackson (2000) on Scots pine (Pinus sylvestris), by
18 Karlsson et al. (2002) on Norway spruce (Picea abies), or by Pelloux et al. (2001) on Aleppo
19 pine (P. halepensis). More recently, Le Thiec and Manninen (2003) reported that drought
20 reduced O3-induced growth suppression of Aleppo pine seedlings. Panek and Goldstein (2001)
21 inferred less impact of O3 on draughted Ponderosa pine (P. ponder osd), and Van den Driessche
22 et al. (1994) reported that drought reduced injury and O3-induced ethylene release by Picea
23 abies. But Karlsson et al. (1997), in a comparative study of fast- and slow-growing clones of
24 P. abies, only observed a drought-induced reduction of O3-inhibited root growth in the fast-
25 growing clone. In contrast, Grulke et al. (2002) noted a synergistic interaction between O3 and
26 drought stress on gross photosynthesis of Pinus ponder osa, and Wallin et al. (2002) reported a
27 synergistic growth response of Picea abies in the third year of a 4-year study. A similar
28 response was noted by Dixon et al. (1998) with the Istebna strain of P. abies.
29 With broad-leaved trees, studies of Durmast oak (Quercuspetraed) (Broadmeadow et al.,
30 1999; Broadmeadow and Jackson, 2000) and European ash (Fraxinus excelsior) (Broadmeadow
31 and Jackson, 2000; Reiner et al., 1996) showed that drought provided partial protection against
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1 O3-induced growth reduction. Although European beech (Fagus sylvaticd) is reportedly an
2 O3-and drought-sensitive species, neither Pearson and Mansfield (1994) nor Broadmeadow et al.
3 (1999) observed any interactions between these stresses, while Dixon et al. (1998) observed
4 partial protection. Paakonen et al. (1998) observed only additive effects in a sensitive clone of
5 birch (Betulapenduld). However, the experiments of Schaub et al. (2003) and the survey by
6 Vollenweider et al. (2003) on black cherry (Prunus serotind) cleary indicate antagonism between
7 drought and O3 stresses on this species.
8 With regard to the converse effect, in a critical review of the evidence for predisposition to
9 drought stress being caused by O3, Maier-Maercker (1998) supported the hypothesis and
10 suggested that the effects were caused by the direct effects of O3 on the walls of the stomatal
11 guard and subsidiary cells in the leaf epidermis, leading to stomatal dysfunction.
12 The Plant Growth Stress Model (PGSW) developed by Chen et al. (1994) is a physiology-
13 based process model which includes drought among several environmental variables.
14 Simulations for Ponderosa pine (Pinus ponder osd) incorporated antagonistic effects between O3
15 and drought stresses, i.e. partial protection, although Karlsson et al. (2000) have since
16 emphasized that drought-induced "memory effects" should be considered when developing
17 simulation models incorporating stomatal conductance.
18 Retzlaff et al. (2000) used the single-tree model, TREEGRO, to simulate the combined
19 effects of O3 and drought on white fir (Abies concolor). Although simulated reductions in
20 precipitation > 25% reduced growth, they also reduced O3 uptake (and impact). But lesser
21 reductions in precipitation combined synergistically with O3 stress to reduce growth, leading the
22 authors to conclude that moderate drought may not ameliorate the response of white fir to O3.
23 On a much larger scale with a modified forest ecosystem model (PnEt-II) incorporating O3-
24 response relationships for hardwood species, Ollinger et al. (1997) showed how predicted
25 changes in net primary production and mean wood production in the northeastern U.S. hardwood
26 forests due to O3 would be reduced (but not countered or reversed) by drought stress, particularly
27 in the southern part of the region. This geographic distribution of the effect was substantiated by
28 the work of Lefohn et al. (1997) on the risk to forest trees in the southern Appalachian
29 Mountains, based on localized estimates of O3 levels and SMD. The TREEGRO and ZELIG
30 models were combined by Laurence et al. (2001) to predict the impacts of O3 and moisture (as
31 precipitation) on the growth of loblolly pine (Pinus taedd) and yellow poplar (Liriodendron
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1 tulipiferd). Based on O3 and precipitation data from three sites in the eastern United States, the
2 six model regressions developed for the two species included both positive and negative
3 coefficients for O3 exposure and precipitation as determinants of growth.
4 As noted in the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996), the effects
5 of soil salinity are similar to those of SMD. In a study of rice (Oryza sativci) cultivars of
6 differing sensitivity to salinity, Welfare et al. (1996) noted that although both O3 and salinity
7 reduced many features of growth additively, antagonistic interactions were only seen for leaf
8 length and potassium accumulation. Similarly, a recent study on chickpea (Cicer arietinum)
9 found no interactions with regard to most components of biomass accumulation (the effects
10 of O3 and salinity were additive), but with root growth, salinity suppressed the adverse effects
11 ofO3.
12 In summary, the recently described interactions of O3 and drought/salinity stresses are
13 consistent with the view that, in many species, drought/salinity reduces the impact of O3 but O3
14 increases sensitivity to drought stress, i.e., the type of response is determined by the sequence of
15 stresses. However, synergisms have also been observed and any antagonisms are species-
16 specific and unpredictable in the absence of experimental evidence. In no case has an
17 antagonism been found to provide complete protection.
18
19 9.4.5 Nutritional Factors
20 The 1996 criteria document (U.S. Environmental Protection Agency, 1996) noted that the
21 large number of macro- and micronutrients and the wide range of species had prevented
22 experimental investigation of all but a few cases of nutrient-O3 interactions and most of these
23 concerned nitrogen (N) and crops or forest tree species. The document also provided a
24 comprehensive tabulation of the results of the relevant studies up to 1992.
25 The suboptimal supply of mineral nutrients to plants leads to various types of growth
26 reductions. The consequences of suboptimal nutrition might, therefore, be expected to have
27 some similarities to those of O3 exposure. One might expect nutritional levels below the
28 optimum either to amplify any effects of O3 or at least lead to additive responses. The difficulty
29 with this suggestion is that the available information has mostly been obtained from
30 experimentation conducted using two or more arbitrarily selected levels of fertility with little or
31 no regard to optima. Hence, it is not surprising that there have been contradictory reports, even
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1 among studies with the same species or cultivars conducted by different workers at different
2 locations using different soils or soil mixes.
3 There appear to have been no recent studies on O3 interactions with specific mineral
4 nutrients other than N. Hence, the previous conclusions are still valid, viz. that increasing levels
5 of the major elements potassium (K) and sulfur (S) usually reduce the impact of O3, or,
6 deficiency increases susceptibility, whereas increased phosphorus (P) usually increases injury,
7 or, deficiency decreases susceptibility.
8 However, with N, a relationship to the optimum is usually demonstrable. Several earlier
9 studies of O3 x N interactions reported that the adverse effects of O3 on growth were greatest at
10 the optimum and decreased with increasing N-deficiency, a finding supported by the work on
11 aspen (Populus tremuloides) of Pell et al. (1995), who also confirmed that excess N decreased O3
12 impact on growth. Similarly, the adverse effects of O3 on growth rate in wheat (Triticum
13 aestivum) diminished with decreased N supply (Cardoso-Vilhena and Barnes, 2001). However,
14 the effects of N are far from consistent. For example, Greitner et al. (1994) reported that O3 and
15 N-deficiency acted additively in aspen in reducing leaf surface area and rate of photosynthesis,
16 Bielenberg et al. (2001) reported that the rate of O3-induced senescence was increased by
17 N-deficiency in hybrid poplar (Populus trichocarpa x maximovizii), and Paakkonen and
18 Holopainen (1995) observed the least adverse effects of O3 on birch (Betulapenduld) at
19 optimum N-fertility levels. With cotton (Gossypium hirsutum), increased N-levels more than
20 overcame the adverse effect of O3 on growth and boll yield (Heagle et al., 1999). In view of
21 these contradictions, one may conclude that other, unrecorded factors may have contributed to
22 the various findings. Thus, much remains unclear about O3 x N-fertility interactions.
23 There have been two recent studies on the effects of overall soil fertility. Whitfield et al.
24 (1998) observed that low general fertility increased O3 sensitivity in selections ofPlantago
25 major. At the biochemical level, well-fertilized birch (Betulapenduld) saplings were found to
26 be less adversely affected by O3 than nutrient-stressed plants (Landolt et al., 1997).
27 TREEGRO model simulations of the growth of red spruce (Picea rubrd) in conditions of
28 nutrient deficiency and O3 stress showed that, in combination, the two stresses acted less than
29 additively (Weinstein and Yanai, 1994). Minimal amelioration by nutrient deficiency was
30 predicted with Ponderosa pine (Pinusponderosa).
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1 Plants may also obtain N and S from airborne sources such as NOX, HNO3, NO3 , SO2, and
2 SO42 , although, depending upon their concentration, these may also be phytotoxic. In various
3 parts of the world, the deposition of N and S in these forms contributes significantly to the levels
4 of nutritionally available N and S in soils. Such depositions may, in turn, influence the impact of
5 O3 on sensitive species through their roles as nutrients independent of any interactions that may
6 occur because of their acidic properties (see Section 9.4.6.5). For example, Takemoto et al.
7 (2001) recently reviewed the situation in southern California's mixed conifer forests and noted
8 that, where N-deposition is appreciable, its combination with O3 is causing a shift in Ponderosa
9 pine (Pinusponderosd) biomass allocation towards that of deciduous trees, with increased
10 needle drop so that only 1- and 2-year needle classes overwinter. Such changes are having
11 significant consequences to the balance of the forest ecosystem and are discussed more fully in
12 Section 9.5.
13 Of the micronutrient elements, only manganese (Mn) appears to have been studied
14 recently. In beans (Phaseolus vulgaris) Mn-deficiency increased O3 toxicity, despite causing
15 reduced O3 uptake (through decreased stomatal conductance) and inducing increased levels of
16 Mn-SOD (Mehlhorn and Wenzel, 1996).
17 In view of the foregoing, it is impossible to generalize about the interactions of soil fertility
18 with O3. While this is especially true of the interactions involving soil nitrogen, for which there
19 is much conflicting evidence, the interactions with other nutrients need much more thorough
20 investigation than has occurred to date, before any clear patterns become apparent.
21
22 9.4.6 Interactions with Other Pollutants
23 The ambient air may be polluted by gases other than O3 and its photochemical oxidant
24 relatives. In particular, industrial, domestic, and automobile emissions and accidents can lead to
25 significant atmospheric concentrations of gases such as sulfur dioxide (SO2), nitric oxide (NO),
26 and nitrogen dioxide (NO2), collectively referred to as NOX, both locally and regionally. Local
27 releases of gases such as hydrogen fluoride (HF), hydrogen chloride (HC1), and chlorine (C12)
28 may result from industrial emissions and accidents. Agricultural fertilizer and manure usage can
29 lead to significant increases in ambient ammonia (NH3). The sulfur and nitrogen oxides may
30 undergo reactions in the atmosphere leading to the formation of sulphate (SO42 ) and nitrate
31 (NO3 ) ions and resultant acid deposition.
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1 The 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996) discounted much of
2 the early research on pollutant combinations, because of its lack of resemblance to the ambient
3 experience: the concentrations used were unrealistically high or the exposure regimes employed
4 almost invariably used gas mixtures, whereas Lefohn et al. (1987) showed that the co-occurrence
5 patterns of significant levels of O3 with SO2 or NO2 in the United States were most frequently
6 sequential or partially sequential with overlap; only rarely were they entirely concurrent. On the
7 other hand, O3 and peroxyacetylnitrate (PAN) frequently co-occur, as both form photochemically
8 under similar conditions.
9 To the list of reviews mentioned in the 1996 criteria document should be added the more
10 recent ones by Barnes and Wellburn (1998), Robinson et al. (1998), and Fangmeier et al. (2002),
11 which also explore some of the potential mechanisms underlying pollutant-pollutant interactions.
12
13 9.4.6.1 Oxidant Mixtures
14 In 1998, Barnes and Wellburn noted that virtually no information existed on the effects on
15 plants of concurrent exposures to O3 and other components of photochemical oxidant other than
16 PAN. The situation has not changed since their review appeared, and the topic appears to have
17 attracted no research interest since before the 1996 O3 AQCD (U.S. Environmental Protection
18 Agency, 1996). The continuing conclusion must, therefore, be that, from the limited information
19 available, the two gases appear to act antagonistically, with O3 raising the threshold for the
20 visible injury response to PAN, and PAN reducing the harmful effects of O3.
21
22 9.4.6.2 Sulfur Dioxide
23 In reviewing O3 x SO2 interactions, Barnes and Wellburn (1998) remarked: "The outcome
24 of exposure to this combination of pollutants has probably been the most studied, yet is one of
25 the least understood." More recent studies have only added to the conflicts referred to in the
26 1996 criteria document (U.S. Environmental Protection Agency, 1996), rather than resolve them.
27 For example, Diaz et al. (1996) reported that, after a year of daily exposures of Aleppo pine
28 (Pinus halepensis)sQQd\ings to 50 ppb O3 and/or 40 ppb SO2, the combination of pollutants
29 synergistically reduced shoot and root growth and impaired mycorrhizal colonization of the
30 roots. With tomato (Lycopersicon esculentum\ on the other hand, effects on growth ranged
31 from synergistic at low exposures (50 ppb) to antagonistic at exposures of 200 ppb of each gas
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1 (Khan and Khan, 1994). Although various physiological measurements were made in these and
2 earlier studies, it has not been possible to determine any consistent mechanism or mechanisms
3 that might account for the conflicting results.
4 Since the information available about O3 x SO2 interactions appears to be highly dependent
5 upon species, the type of response measured, and the experimental protocol used, it would still
6 appear prudent to heed the statement of Heagle et al. (1988) in their summary of the studies
7 undertaken in 12 field experiments over several years within the NCLAN program: "There were
8 no cases where O3 and SO2 interactions significantly affected yield." (emphasis added.)
9
10 9.4.6.3 Nitrogen Oxides, Nitric Acid Vapor, and Ammonia
11 The major oxides of nitrogen that occur in ambient air are nitrous oxide (N2O), nitric oxide
12 (NO) and nitrogen dioxide (NO2), of which the latter two (conveniently symbolized as NOX) are
13 particularly important in connection with O3, because they are components of the reaction mix
14 that leads to photochemical O3 formation and because they can interact with O3-responses. Their
15 reactions in the atmosphere can also lead to the occurrence of nitric acid vapor (HNO3) in
16 ambient air. The other major N-containing contaminant of ambient air in many parts of the
17 world is ammonia (NH3), largely released through agricultural practices.
18 Despite various combinations of O3 and NOX being probably the most common air
19 pollutant combinations found in the field, Barnes and Wellburn (1998) noted that they have been
20 little studied. Much early work with O3 and NOX focused on O3 x NO2 interactions and can be
21 discounted, because of the unrealistic concentrations employed and their use as mixtures rather
22 than in types of sequences. The 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996)
23 concluded that evidence from studies involving concurrent exposures to both O3 and NOX at
24 realistic concentrations was so fragmented and varied that no firm conclusions could be drawn as
25 to the likelihood and nature such interactions. However, the few recent investigations taken
26 together with the earlier data are now beginning to reveal a pattern of response.
27 With regard to NO, Nussbaum et al. (1995a; 2000b) reported their findings with concurrent
28 exposures to NO and O3 and observed that, at low O3 levels, NO tended to act similarly to O3 by
29 increasing the scale of responses such as growth reductions. However, in ambient air in which
30 O3 is a dominant factor, the effects of NO were usually found to be negligible due to low levels,
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1 although the authors admitted that the effects observed were confounded by the inevitable
2 O3-induced oxidation of NO to NO2.
3 Two possible mechanisms whereby NO may influence plant response to O3 are suggested
4 by recent biochemical studies. First, there is growing evidence for the role of NO as a signaling
5 agent in plants that can induce defense responses to a range of biotic and abiotic stressors
6 (Beligni and Lamattina, 2001; Neill et al., 2002). Second, a role for NO as an antioxidant
7 scavenger of reactive oxygen species has been demonstrated by Beligni and Lamattina (2002) in
8 potato (Solarium tuberosum) leaves and chloroplasts. However, both of these cases concern
9 endogenously synthesized NO, and it must be noted that in none of these or other reports of
10 studies of NO signaling have the authors considered the potential significance of exogenous NO
11 in ambient air.
12 An independent case for O3 x NO interactions comes from Mills et al. (2000). The ANN
13 model developed to predict the O3 effects on white clover (Trifolium repens) biomass based on
14 experiments at 18 locations throughout Europe found that the minimum daily NO concentration
15 (at 5 p.m.) was a significant contributor to adverse effects.
16 Turning to NO2, Maggs and Ashmore (1998) found that, although concurrent but
17 intermittent exposures of Bismati rice (Oryza saliva) revealed no significant growth interactions,
18 NO2 reduced the rate of O3-induced senescence, an antagonistic response possibly related to
19 enhanced N-metabolism.
20 With regard to sequential exposures, two studies on gene activation in tobacco (Nicotiana
21 tabacum) revealed that NO2 counteracted the effect of O3 in reducing mRNA levels for three
22 genes encoding photosynthetic proteins (Bahl and Kahl, 1995) and tended to counteract the
23 O3-induced enhancement of defense-protein gene activation (Bahl etal., 1995). However,
24 despite compelling evidence for significant interactive effects provided by earlier studies
25 (Bender et al., 1991; Goodyear and Ormrod, 1988; Runeckles and Palmer, 1987), the only recent
26 investigation of growth effects seems to have been that of Mazarura (1997) using sine-wave
27 exposure profiles. He found that although 4 weeks of twice daily 3-h exposures to NO2 (120 ppb
28 peak concentrations) slightly stimulated growth of radish (Raphanus saliva] and while daily 6-h
29 exposures to O3 (120 ppb peak concentration) did not significantly reduce growth, the daily
30 sequence, NO2 - O3 - NO2, led to a 13% drop in dry matter production.
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1 The combined evidence to date, therefore, suggests that, in leguminous species, the effects
2 of these sequences are antagonistic with NO2 tending to reduce (or reverse) the negative effects
3 of O3 on growth, while the effects are increased in other species. These conclusions differ from
4 those of Barnes and Wellburn (1998) who suggested that sequential exposures tended to result in
5 antagonistic effects (largely based on the summary by Bender and Weigel, 1992), whereas
6 simultaneous exposures were likely to lead to synergistic responses. With disagreements both
7 among the data and their interpretation, it is not possible to determine the circumstances under
8 which specific interactions of O3 and NO2 may occur, but there is no reason to doubt the validity
9 of the individual findings of each study. Far more systematic investigation is needed to clarify
10 the situation.
11 There appear to have been no studies of O3 interacting with HNO3 in the vapor phase.
12 However, in the southern California montane forests (Takemoto et al., 2001), in Sweden (Janson
13 and Granat, 1999), and elsewhere, significant amounts of N are deposited in this form because of
14 the vapor's high deposition velocity. As a consequence, although much of it ultimately reaches
15 the ground through leaching and leaf fall and enters the soil as nitrate (NO3 ), it may also be used
16 as a N source by the foliage itself. This nutritional role is independent of any contribution that
17 HNO3 vapor may make to acidic deposition. Indirect interactions with the effects of O3 through
18 N-deposition of NOX, HNO3, and NH3 are related to the interactions of O3 with N as a nutrient,
19 and have recently been examined in the review by Takemoto et al. (2001). The 1996 O3 AQCD
20 (U.S. Environmental Protection Agency, 1996) stated that the evidence available at that time led
21 to estimates of total forest dry deposition, including HNO3, ranging from 5.7 to 19.1 kg N ha"1
22 year1 (Taylor, Jr. et al., 1988). However, Takemoto et al. (2001) pointed out that in parts of the
23 mid-elevation forests of southern California, dry deposition rates may reach more than 40 kg N
24 ha"1 year1. As a result, some locations have seen the conversion from N-limited to N-saturated
25 forests. The concern for California's forests is well-stated by Takemoto et al.: "As potential
26 modifiers of long-term forest health, O3 is a stressor and N deposition is an enhancer of
27 ponderosa/Jeffrey pine physiology and growth (Grulke and Balduman, 1999). The progression
28 toward a deciduous growth habit, higher shootroot biomass ratios, increasing depths of litter,
29 tree densification, and elevated NO3- levels in soil and soil solution, all point to the replacement
30 of pine species with nitrophilous, shade- and O3-tolerant tree species, such as fir and cedar
31 (Minnich et al., 1995) (Minnich, 1999)."
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1 Few studies have been reported of interactions of O3 with NH3. The 1996 criteria
2 document made reference to the work on bean (Phaseolus vulgaris) by Tonneijck and van Dijk
3 (1994, 1998). Although NH3 alone tended to increase growth and O3 alone to inhibit it, one
4 interaction was noted (Tonneijck and van Dijk, 1994) on the number of injured leaves. Dueck
5 et al. (1998) studied the effects of O3 and NH3 on the growth and drought resistance of Scots
6 pine (Pinus sylvestris). Significant interactions were found for some growth features, but there
7 were no consistent patterns of the effects of NH3 on O3 response or vice versa. However, O3 was
8 found to ameliorate the enhancement of drought stress caused by NH3.
9 At this time there is insufficient information to offer any general conclusions about the
10 interactive effects of O3 and NH3.
11
12 9.4.6.4 Hydrogen Fluoride and Other Gaseous Pollutants
13 Although HF and other fluorides are important local air pollutants associated with
14 aluminum smelting and superphosphate fertilizer manufacture, no studies of possible interactions
15 with oxidants appear to have been reported since that of MacLean (1990). He found that HF
16 retarded the accelerated senescence and loss of chlorophyll resulting from O3 exposure in corn
17 seedlings. However, such an isolated observation cannot be taken to indicate that HF can reduce
18 the impact of O3 on other species or even that the effect would ultimately have led to an effect on
19 mature plants.
20
21 9.4.6.5 Acid Deposition
22 The deposition of acidic species onto vegetation may elicit direct effects on the foliage or
23 indirect effects via changes induced in the soil. The 1996 O3 AQCD (U.S. Environmental
24 Protection Agency, 1996) included an extensive listing of investigations into the effects of O3
25 and acid deposition (usually in the form of simulated acid rain, SAR) on plant growth and
26 physiology. The majority of studies found no effects of SAR or acid mists or fogs at pH values
27 greater than about 3.0 and no interactive effects with O3. (In ambient air, pH values less than 3.0
28 have rarely been reported.) In the few reports in which significant interactions were found, most
29 were antagonistic and were explained as probably being the result of increased fertility due to
30 nitrate and sulfate supplied in the SAR.
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1 Although numerous reviews have recently appeared (e.g., Bussotti and Ferretti, 1998;
2 Fliickiger et al., 2002; Fox and Mickler, 1996; Nussbaum et al., 1999; and Sheppard and Cape,
3 1999), the shift in interest in air pollution effects away from acid deposition has resulted in little
4 new research having been reported over the past 10 or so years. In most of the reported studies,
5 no effects due to the O3 exposures, the SAR treatments used, or their combinations were
6 observed, e.g., Baker et al. (1994) on loblolly pine (Pinus taeda); Laurence et al. (1997), and
7 Vann et al. (1995) on red spruce (Picea rubens); and Laurence et al. (1996) on sugar maple
8 (Acer saccharum). Branch chamber studies of 12-year old Ponderosa pine (Pinusponderosd)
9 trees by Momen et al. (1997, 1999) revealed no O3 effects or interactions. With red spruce
10 (Picea rubens), Sayre and Fahey (1999) noted no effects of O3 on the foliar leaching of Ca or
11 Mg, which only became significant with SAR at pH 3.1. Izuta (1998) observed no interactions
12 with Nikko fir (Abies homolepis), although SAR at ph 4.0 reduced dry matter. Shan et al. (1996)
13 reported adverse effects of O3 but none attributable to SAR on the growth of Pinus armandi.
14 With herbaceous species, Ashenden et al. (1995) noted significant antagonistic interactions
15 of O3 and acid mist in white clover (Trifolium repens\ in which the adverse effect of low pH was
16 countered by O3. In contrast, Ashenden et al. (1996) found that, although pH 2.5 mist caused a
17 significant stimulation of the growth of ryegrass (Lolium perenne) attributed to a fertilizer effect,
18 and O3 caused reduced growth, there was no interaction. Bentgrass (Agrostis capillaris) behaved
19 similarly.
20 A study by Bosley et al. (1998) on the germination of spores of the moss, Polytrichum
21 commune, and the ferns, Athyrium felix-femina and Onoclea sensibilis, revealed no effect of O3
22 on moss spores, while SAR at pH < 4.0 was completely inhibitory. With the ferns, germination
23 was progressively reduced by both increased O3 and acidity.
24 In summary, the few findings of interactions in these recent studies are consistent with the
25 previous conclusion regarding the likelihood of such interactions being antagonistic. However,
26 the interactions observed were in each case largely the result of the response to the lowest pH
27 used, which in several cases was below 3.0, and hence may not be relevant to most field
28 conditions.
29
30
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1 9.4.6.6 Heavy Metals
2 Since there appears to have been no further research into the interactions of oxidants with
3 heavy metal pollutants, our understanding is unchanged from at the time of the 1996 O3 AQCD
4 (U.S. Environmental Protection Agency, 1996). As noted therein, the limited data available
5 from early studies indicates varying degrees of enhancement of any adverse effects of O3 but
6 precludes the development of any response relationships.
7
8 9.4.6.7 Mixtures of Ozone with Two or More Pollutants
9 In many airsheds the mixtures that occur, both concurrently and over time, may involve
10 three or more pollutants. Very little useful information exists on the effects of O3 with multiple
11 pollutants. As the 1996 criteria document and others have pointed out, most of the early studies
12 on such combinations can be discounted because of their use of (1) high and environmentally
13 irrelevant exposure concentrations and (2) unrealistic, repetitive exposure profiles (U.S.
14 Environmental Protection Agency, 1996; Barnes and Wellburn, 1998).
15 The large investment in experimental facilities required to study these complex interactions
16 is a major deterrent. So, although the topic has been included in several reviews that have
17 appeared in the last decade, there appear to have been only two studies that have provided new
18 information on the effects of O3 in combination with more than one other pollutant stress.
19 Ashenden et al. (1995, 1996) studied the effects of O3 and/or (SO2 + NO2) with four acidities of
20 SAR applied to each gas treatment, on white clover (Trifolium repens) and two pasture grasses
21 (Lolium perenne andAgrostis capillaris). With each species, the antagonism reported for the O3
22 x SAR interaction (Section 9.4.4.6.5) tended to be nullified by concurrent exposure to the other
23 gases, while the combination of the three gaseous pollutants resulted in the most severe growth
24 inhibition, regardless of the acidity of the SAR.
25 With such meager evidence, no clear conclusions can be drawn as to the ways in which the
26 effects of multiple airborne stressors could influence or be influenced by O3.
27
28 9.4.7 Interactions with Agricultural Chemicals
29 The review of interactions involving O3, plants, and various agricultural chemicals
30 presented in the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996) remains a valid
31 assessment of our limited knowledge of these interrelationships. Our knowledge is largely based
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1 on the protection against O3 afforded to a range of crop species by applications of various
2 chemicals, particularly fungicides, such as benomyl (benlate; methyl-l-[butylcarbamoyl]-2-
3 benzimidazolecarbamate) and several carbamates and triazoles. A recent report has added
4 azoxystrobin (AZO) and epoxyconazole (EPO) to the list (Wu and Tiedemann, 2002). Foliar
5 sprays of either AZO or EPO provided 50 to 60% protection against O3 injury to barley
6 (Hordeum vulgare) leaves. Both had similar modes of action involving stimulation of the levels
7 of antioxidant enzymes such as SOD, ascorbate peroxidase, guaiacol peroxidase, and catalase.
8 In contrast, applications of herbicides have yielded variable results ranging from increased
9 sensitivity to protection from O3; the nature of the effect is usually species- or cultivar-
10 dependent. Although of less wide application, some plant growth retardants have also been
11 found to provide protection, but no insecticide appears to have been clearly shown to have
12 similar properties.
13 Despite the attraction of the use of permitted chemicals to provide crop protection, the
14 statement in the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996) is still valid:
15 "It is premature to recommend their use specifically for protecting crops from the adverse effects
16 of O3, rather than for their primary purpose."
17
18 9.4.8 Factors Associated with Global Climate Change
19 During the last decade, interest in the effects of climatic change on vegetation has replaced
20 concerns over the purported causes of forest decline and the effects of acidic deposition. Two
21 specific components of climate change have been singled out as the foci of most of the research
22 activity:
23 • the effects of increasing mean global CO2 concentrations in the lower atmosphere, and
24 • the effects of increasing levels of surface-level irradiation by UV-B (the result of
stratospheric O3 depletion).
25 In spite of the crucial role of temperature as a climatic determinant (Monteith and Elston, 1993),
26 the effects of increasing mean global temperatures and their interactions with increasing CO2
27 levels in particular have received less attention.
28 All of the biotic and chemical interactions with oxidants discussed in the preceding
29 sections may be modified by these climatic changes. However, research activities have largely
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1 focused on the two-way O3 x CO2 interaction. Little if any experimental evidence exists related
2 to three-way interactions such as O3 x CO2 x disease or O3 x CO2 x nutrient availability,
3 although such interactions cannot be predicted from the component two-way interactions.
4 Numerous reviews have appeared since the 1996 O3 AQCD (U.S. Environmental
5 Protection Agency, 1996) dealing with the issues involved. General reviews include
6 publications of IPCC (1996); IPCC (2001); and UNEP (1993, 1999); the volume by Wellburn
7 (1994); the volumes edited by Alscher and Wellburn (1994), De Kok and Stulen (1998), Singh
8 (2000), and Yunus and Iqbal (1996); and papers by Idso and Idso (1994), Krupa and Groth
9 (2000), Luo et al. (1999), Polle and Pell (1999), Poorter and Perez-Soba (2001), Runeckles
10 (2002), and Weubbles et al. (1999). Effects on agriculture and crop production, growth, and
11 metabolism have been reviewed by Groth and Krupa (2000), Rotter et al. (1999), and Schnug
12 (1998); effects on forests have been reviewed by Bortier et al. (2000); with focus on insect pests,
13 Docherty et al. (1997), Karnosky et al. (2001b,a), McLaughlin and Percy (1999), and Saxe et al.
14 (1998).
15 As background to the discussion of interactions with O3, it should be noted that the
16 increased levels of CO2 experienced since the mid-18th century are such that, without abatement
17 of the rates of increase, increased levels of from 540 to 970 ppm have been projected by the year
18 2100 (IPCC, 2001). Such increases in the concentration of CO2, the principal GHG released into
19 the atmosphere, will inevitably lead to increased global mean temperatures, evidence for which
20 is already available from oceanic, icepack, and other records. The latest estimates of the global
21 warming are for an increase in the range 1.4 to 5.8 °C over this century, in contrast to the 0.6 °C
22 rise experienced since 1900 (IPCC, 2001). However, considerable uncertainty is associated with
23 such projections of future increases in global temperature.
24 The use of elevated CO2 concentrations has been common practice for many years in the
25 production of many greenhouse crops. Much of our early knowledge of the effects of higher
26 than ambient CO2 levels on plant growth derives from this application, coupled with research of
27 plant physiologists on how CO2 concentrations affect the process of photosynthesis. Information
28 available about effects of increased CO2 levels on photosynthesis and stomatal function, in
29 particular, has provided the underlying bases for numerous process models that simulate plant
30 growth under stress and in changed climates.
January 2005 9-127 DRAFT-DO NOT QUOTE OR CITE
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1 Although simple O3 x temperature interactions were discussed in Section 9.4.4.2, the close
2 linkage between global CO2 levels and global mean temperatures in the context of climate
3 change requires that an assessment of the interactive effects with O3 should focus, as much as
4 possible, on interactions involving all three factors.
5
6 9.4.8.1 Ozone-Carbon Dioxide-Temperature Interactions
7 Idso and Idso (1994) reviewed several hundred reports published between 1982 and 1994
8 on the effects of increased CO2 on plant growth and net photosynthesis. Their survey covered a
9 wide range of temperate and tropical, herbaceous and perennial species, including coniferous
10 trees. They concluded that, for responses to a 300-ppm increase in CO2, somewhat less than a
11 doubling of present-day levels, but somewhat greater than the 540 ppm lower limit suggested by
12 the IPCC (2001), averaged across all species:
13 • light intensity had a negligible effect on net photosynthesis other than at limiting low
intensities under which the CO2-driven enhancement was increased;
14 • increased temperature tended to increase the CO2-driven enhancement of dry matter
accumulation (growth) and net photosynthesis;
15 • drought conditions tended to increase the CO2-driven enhancements of both growth and
net photosynthesis, but increased salinity had little effect;
16 • mineral nutrient deficiency (especially of nitrogen) tended to increase the CO2-driven
enhancement of growth; and
17 • in the presence of air pollutants (especially SO2 and NOX), the CO2-driven enhancement
of net photosynthesis tended to be increased.
18 It should be noted that the statement that CO2-enhanced growth increased with temperature
19 referred to total dry matter accumulation by the whole plant and not to the yield of grain, fruit, or
20 seed. Unfortunately, despite the existence of several reports at the time, the summary of
21 interactions with air pollutants contained only a single reference to O3, i.e., Pfirrmann and
22 Barnes (1993), who reported surprisingly that a doubling of CO2 levels led to a 27% increase in
23 dry weight of radish (Raphanus sativus) but that the combination with O3 led to a 77% increase.
24 The more recent reviews by Rudorff et al. (2000) and Olszyk et al. (2000) have addressed
25 CO2 x O3 interactions in detail, with the latter focusing on the implications for ecosystems. They
26 concluded that:
January 2005 9-128 DRAFT-DO NOT QUOTE OR CITE
-------
1 • the effects of both gases on stomatal closure were predominantly additive, with little
evidence of interaction;
2 • increased photosynthesis resulting from elevated CO2 may be canceled by exposures to
high O3 levels;
3 • foliar O3 injury is reduced by elevated CO2; and
4 • interactions between CO2 and O3 can affect storage carbohydrates, leaf free-radical
metabolism, and carbon allocation to shoots and roots.
5 Olszyk et al. also made specific note of the relative lack of information on below-ground effects.
6 Much of the recently published information on the effects of increased CO2 and O3 levels is
7 summarized in Table 9-12. Note that the table only lists the directions of O3-induced effects and
8 any modifications of these effects resulting from elevated CO2, not their magnitudes. These
9 directions are usually, but not necessarily, the same as the corollary effects of O3 on
10 CO2-induced responses.
11 The bulk of the available evidence clearly shows that, under the various experimental
12 conditions used (which almost exclusively employed abrupt or "step" increases in CO2
13 concentration, as discussed below), increased CO2 levels may protect plants from the adverse
14 effects of O3 on growth. This protection may be afforded in part by CO2 acting together with O3
15 in inducing stomatal closure, thereby reducing O3 uptake, and in part by CO2 reducing the
16 negative effects of O3 on Rubisco and its activity in CO2-fixation. However, the situation with
17 regard to the combined effects of O3 and CO2 on stomatal functioning is not clear-cut. In spite of
18 the wealth of evidence supporting both CO2-induced and O3-induced short-term stomatal closure,
19 studies over long-term exposures, especially with tree species, have revealed little effect of
20 elevated CO2 on stomatal conductance. Reverse effects of O3 on stomata have also been noted,
21 as a result of O3-induced stomatal dysfunction after extended periods of exposure (Maier-
22 Maercker, 1998).
23 At the mechanistic level, Rubisco plays a key role in CO2-assimilation, and while both O3
24 and elevated CO2 per se can lead to reduced activity, CO2 can also reverse the O3-induced
25 inhibition of Rubisco activity and photosynthesis (Table 9-12). However, in their review of the
26 possible mechanisms involved, Polle and Pell (1999) cautioned that Rubisco should not be
27 regarded "as a unique target for the interaction of the two gases." But it is clear from the bulk of
28
January 2005 9-129 DRAFT-DO NOT QUOTE OR CITE
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3
Table 9-12. Effects of Increased Carbon Dioxide on Ozone-Induced Responses of Plants at the Metabolic,
Physiological, and Whole-Plant Levels
to
o
o
VO
1
o
o
H
6
O
0
H
O
O
0
o
H
W
Co2 Effects:
Plant response O3 Response3
Biochemical/metabolic
Ascorbate peroxidase V
V
V
Catalase V
A
Chlorophyll V
V
Glutathione reductase V
O
O
Rubisco V
V
[A]
V
O3 Effects: V, Decrease; A, Increase; O, No Significant Effect.
A, Additive/Synergistic; T, Antagonistic/Ameliorative; O, No Significant Effect.
CO2 Modification"
V
OD
OD
OD
OD
T
V
T
OD
OD
T
V
V
T
Species
Wheat (Triticum aestivum)
Sugar maple (Acer saccharum)
Trembling aspen
(Populus tremuloides)
Wheat (T. aestivum)
Wheat (T. aestivum)
Potato (Solanum tuberosum)
Wheat (T. aestivum)
Sugar maple (A. saccharum)
Aspen (P. tremuloides)
Soybean (Glycine max)
Wheat (T. aestivum)
Trembling aspen
(P. tremuloides)
Sugar maple (A. saccharum)
Facility0
CSTR, P
CEQP
FACE, G
CEQP
CEQP
OTC, G
OTC, G
CSTR, P
CEQP
FACE, G
OTQP
OTC, G
FACE, G
CEQP
Reference
Rao etal. (1995)
Niewiadomska et al. (1999)
Wustman etal. (2001)
McKeeetal. (1997b)
Niewiadomska et al. (1999)
Ommenetal. (1999);
Donnelly et al. (2000)
Donnelly et al. (200 la)
Rao etal. (1995)
Niewiadomska et al. (1999)
Wustman etal. (2001)
Reid etal. (1998)
McKee et al. (2000)
Noormets et al. (2001)
Gaucher et al. (2003)
-------
3
Table 9-12 (cont'd). Effects of Increased Carbon Dioxide on Ozone-Induced Responses of Plants at the
Metabolic, Physiological, and Whole-Plant Levels
to
o
o
VO
1
o
H
6
O
0
H
O
O
0
o
H
W
O3 Effects: V, Decrease; A, Increase; O, No Significant Effect.
Co2 Effects: A, Additive/Synergistic; T, Antagonistic/Ameliorative; O, No Significant Effect.
Plant response O3 Response3 CO2 Modification6
Biochemical/metabolic (cont'd)
Rubisco activity V V
V T
V [T]
V T
Superoxide dismutase O OD
O OD
Physiological
Stomatal conductance V A
V T
V T
V [A]
A - OD T - OD
V A
Species
Soybean (G. max)
Wheat (T. aestivum)
Wheat (T. aestivum)
European beech
(Fagus sylvatica)
Wheat (T. aestivum)
Sugar maple (A. saccharum)
Radish (Raphanus sativus)
Soybean (G. max)
Bean (Phaseolus vulgaris)
White clover (Trifolium
repens) (O3-sensitive)
White clover (T. repens)
(O3-tolerant)
Tomato (Lycopersicon
esculentum)
Facility0
OTC,P
CEC,P
OTC, G
CEC,P
CEC,P
CEC,P
CEC,P
OTC, G
OTC,P
CSTR, P
CSTR, P
CEC,P
Reference
Reid etal. (1998)
McKeeetal. (1995)
McKee et al. (2000)
Liitz et al. (2000)
McKeeetal. (1997b)
Niewiadomska et al. (1999)
Barnes and Pfirrmann (1992)
Mulchi etal. (1992)
Heagle et al. (2002)
Heagle etal. (1993)
Heagle etal. (1993)
Hao et al. (2000)
-------
3
Table 9-12 (cont'd). Effects of Increased Carbon Dioxide on Ozone-Induced Responses of Plants at the Metabolic,
Physiological, and Whole-Plant Levels
to
o
o
VO
1
to
o
H
1
O
o
o
H
O
o
H
w
o
^^
o
H
W
O3 Effects: V, Decrease; A, Increase; O, No Significant Effect.
Co2 Effects:
Plant response O3 response3
Physiological (cont'd)
Stomatal conductance V
(cont'd)
V
[V]
V
V
A
V
V
OD
OD
OD
OD
[A]
V
: A, Additive/Synergistic; T, Antagonistic/Ameliorative; O, No Significant Effect.
CO2 modification*
A
A
A
OD
O-A
OD
OD
OD
OD
OD
OD
OD
T
A
Species
Potato (S, tuberosum)
Wheat (T. aestivum)
Agropyron smithii
Koeleria cristata
Bouteloua curtipendula
Schizachyrium scoparium
Black cherry (Prunus serotina)
Green ash
(Fraxinus pennsylvanica)
Yellow poplar
(Liriodendron tulipiferd)
Trembling aspen
(P. tremuloides)
Facility0
OTC, G
CEQP
CEQP
OTC, G
CEQP
CEQP
CEQP
CEQP
CEQP
CEQP
CSTR, P
CSTR, P
CSTR, P
CEQP
Reference
Finnan et al. (2002)
Balaguer etal. (1995)
Barnes etal. (1995)
McKee etal. (1995)
Mulholland et al. (1997b)
Donnelly etal. (1998)
Tiedemann and Firsching (2000)
Volin etal. (1998)
Volin etal. (1998)
Volin etal. (1998)
Volin etal. (1998)
Loats and Rebbeck (1999)
Loats and Rebbeck (1999)
Loats and Rebbeck (1999)
Volin etal. (1998)
-------
3
Table 9-12 (cont'd). Effects of Increased Carbon Dioxide on Ozone-Induced Responses of Plants at the Metabolic,
Physiological, and Whole-Plant Levels
to
o
o
VO
1
OJ
O
s
H
6
o
0
H
O
o
0
o
H
W
Co2 Effects:
Plant response O3 Response3
Physiological (cont'd)
Stomatal conductance O - V
(cont'd)
OD
[A]
Photosynthesis V
V
V
v
V
OD
OD
[V]
V
V
O3 Effects: V, Decrease; A, Increase; O, No Significant Effect.
A, Additive/Synergistic; T, Antagonistic/Ameliorative; O, No Significant Effect.
CO2 Codification"
V
OD
V
V
T
T
T
OD
OD
[T]
V
V
OD
Species
Red oak (Quercus rubra)
Durmast oak (Quercus petraed)
Radish (R. sativus)
Soybean (G. max)
Soybean (G. max)
Bean (P. vulgaris)
Tomato (L. esculentum)
Potato (S. tuberosum)
Wheat (T. aestivum)
Wheat (T. aestivum)
Facility0
FACE, G
CEQP
CEQP
CEQP
OTQP
OTQG
OTQP
CEQP
OTQG
OTQG
CEQP
OTQG
OTQP
CEQP
CEQP
Reference
Noormets et al. (2001)
Volin etal. (1998)
Broadmeadow et al. (1999)
Barnes and Pfirrmann (1992)
Booker etal. (1997)
Mulchi etal. (1992)
Heagle et al. (2002)
Hao et al. (2000)
Donnelly et al. (200 la)
Lawson etal. (200 Ib)
Barnes etal. (1995)
Donnelly et al. (2000)
Mulholland et al. (1997b)
Reid and Fiscus (1998)
Tiedemann and Firsching (2000)
Cardoso-Vilhena and Barnes (2001)
-------
3
Table 9-12 (cont'd). Effects of Increased Carbon Dioxide on Ozone-Induced Responses of Plants at the Metabolic,
Physiological, and Whole-Plant Levels
to
o
o
VO
1
o
H
6
O
0
H
/->
Co2 Effects:
Plant response O3 Response3
Physiological (cont'd)
Photosynthesis (cont.) OD
V
o-v
[V]
V
V
OD
[V]
OD
V
V
OD
V
O3 Effects: V, Decrease; A, Increase; O, No Significant Effect.
A, Additive/Synergistic; T, Antagonistic/Ameliorative; O, No Significant Effect.
CO2 Codification"
OD
T
OD
T
T
T
OD
T
OD
T
T
OD
T
Species
Agropyron smithii
Koeleria cristata
Bouteloua curtipendula
Schizachyrium scoparium
Ponderosa pine
(Pinus ponderosa)
Scots pine (Pinus sylvestris)
Black cherry (P. serotina)
Green ash (F. pennsylvanica)
Yellow poplar (L. tulipifera)
Trembling aspen
(P. tremuloides)
European beech (F. sylvatica)
Red oak (Q. rubra)
Sugar maple (A. saccharum)
Facility0
CEQP
CEQP
CEQP
CEQP
CEQG
OTQ G
CSTR, P
CSTR, P
CSTR, P
CEQP
CEQP
CEQP
CEQP
Reference
Volinetal. (1998)
Volinetal. (1998)
Volinetal. (1998)
Volinetal. (1998)
Olszyketal. (2001)
Kellomaki and Wang (1997b);
Kellomaki and Wang (1997a)
Loats and Rebbeck (1999)
Loats and Rebbeck (1999)
Loats and Rebbeck (1999)
Volinetal. (1998)
Grams etal. (1999)
Volinetal. (1998)
Gaucher et al. (2003)
O
HH
H
W
-------
3
Table 9-12 (cont'd). Effects of Increased Carbon Dioxide on Ozone-Induced Responses of Plants at the Metabolic,
Physiological, and Whole-Plant Levels
to
o
o
VO
1
0,
O
H
6
o
0
H
O
O
s
0
o
H
W
O3 Effects: V, Decrease; A, Increase; O, No Significant Effect.
Co2 Effects: A, Additive/Synergistic; T, Antagonistic/Ameliorative; O, No Significant Effect.
Plant response O3 Response3 CO2 Codification"
Physiological (cont'd)
Photorespiration V OD
V A
Growth, Yield
Total biomass V V
V T
V T
V T
[V] OD
V T
OD OD
V T
OD OD
Species
Soybean (G. max)
Wheat (T. aestivum)
Parsley (Petroselinum sativum)
Bean (P. vulgaris)
Soybean (G. max)
Alfalfa (Medicago sativa)
White clover (T. repens)
(O3-sensitive)
White clover (T. repens)
(O3-tolerant)
Tomato (L. esculentum)
Potato (S. tuberosum)
Facility0
OTQP
CEQP
CEQP
CEQP
OTQP
OTQG
CSTR, P
CEQP
CSTR, P
CSTR, P
CEQP
CSTR, P
OTQG
Reference
Booker etal. (1997)
McKeeetal. (1997b)
Cardoso-Vilhena et al. (1998)
Cardoso-Vilhena et al. (1998)
Heagle et al. (2002)
Mulchi etal. (1992)
Reinertetal. (1997)
Johnson etal. (1996a)
Heagle etal. (1993)
Heagle etal. (1993)
Hao et al. (2000)
ReinertandHo(1995)
Donnelly etal. (200 Ib);
Persson et al. (2003)
-------
3
Table 9-12 (cont'd). Effects of Increased Carbon Dioxide on Ozone-Induced Responses of Plants at the Metabolic,
Physiological, and Whole-Plant Levels
to
o
o
VO
1
OJ
Oi
O
f?
H
1
O
o
0
H
O
o
0
o
H
W
O3 Effects: V, Decrease; A, Increase; O, No Significant Effect.
Co2 Effects:
Plant response O3 Response3
Growth, Yield (cont'd)
Total biomass (cont'd) V
V
V
v
V
[V]
V
V
[V]
V
V
A, Additive/Synergistic; T, Antagonistic/Ameliorative; O, No Significant Effect.
CO2 Codification"
0
T
T
T
T
V
OD
[V]
V
V
V
Species
Mustard (Sinapis alba)
Plantain (Plantago major)
Cotton (Gossypium hirsutum)
Wheat (T. aestivum)
Wheat (T. aestivum)
Wheat (T. aestivum)
Wheat (T. aestivum)
Timothy (Phleum pratense)
Agropyron smithii
Koeleria cristata
Facility0
OTC, G
CEC,P
CEC,P
OTC,P
CEC,P
OTC, G
OTC,P
CEC,P
OTC, G
CSTR, P
OTC, G
OTC, G
CEC,P
OTC, G
CEC,P
CEC,P
CEC,P
Reference
Lawsonetal. (200 la)
Cardoso-Vilhena et al. (1998)
Cardoso-Vilhena et al. (1998)
Booker, 2000)
Heagle etal. (1999)
Cardoso-Vilhena et al. (1998)
Fangmeier et al. (1996)
Heagle et al. (2000)
McKeeetal. (1997a)
Pleijel et al. (2000)
Rao etal. (1995)
Rudorffetal. (1996a)
Bender etal. (1999);
Mulholland et al. (1997a)
Cardoso-Vilhena et al. (1998)
Tiedemann and Firsching (2000)
Ewart and Pleijel (1999)
Johnson etal. (1996a)
Volin etal. (1998)
Volin etal. (1998)
-------
3
Table 9-12 (cont'd). Effects of Increased Carbon Dioxide on Ozone-Induced Responses of Plants at the Metabolic,
Physiological, and Whole-Plant Levels
to
o
o
VO
1
3
o
^
H
6
O
0
H
O
O
s
0
o
H
W
Co2 Effects:
Plant response O3 Response3
Growth, Yield (cont'd)
Total Biomass (cont'd) V
V
O
V
V
[V]
A
O
[A]
V
V
V
0
V
O3 Effects: V, Decrease; A, Increase; O, No Significant Effect.
A, Additive/Synergistic; T, Antagonistic/Ameliorative; O, No Significant Effect.
CO2 Codification"
V
OD
OD
T
A
OD
V
OD
A
V
V
OD
OD
T
Species
Corn (Zea mays)
Bouteloua curtipendula
Schizachyrium scoparium
Ponderosa pine (P. ponderosa)
Birch (Betula pendula)
Black cherry P. serotina
Green ash (F. pennsylvanicd)
European ash
(Fraxinus excelsior)
Yellow poplar (L. tulipifera)
Sugar maple (A. saccharum)
Trembling aspen
(P. tremuloides)
(O3-tolerant clone}
(O3-sensitive clone)
Red oak (Q. rubra)
Durmast oak (Q. petraea)
Facility0
OTC, G
CEC,P
CEC,P
CEC, G
CEC,P
CSTR, P
CSTR, P
OTC, G
CSTR, P
CEC,P
CEC,P
OTC,P
OTC, G
OTC, G
CEC,P
CEC,P
OTC, G
Reference
Rudorffetal. (1996a)
Volinetal. (1998)
Volinetal. (1998)
Olszyketal. (2001)
Kytoviita et al. (1999)
Loats and Rebbeck (1999)
Loats and Rebbeck (1999)
Broadmeadow and Jackson (2000)
Loats and Rebbeck (1999)
Gaucher et al. (2003)
Volinetal. (1998)
Dicksonetal. (1998)
Dicksonetal. (2001)
Dicksonetal. (2001)
Volinetal. (1998)
Broadmeadow et al. (1999)
Broadmeadow and Jackson (2000)
-------
Table 9-12 (cont'd). Effects of Increased Carbon Dioxide on Ozone-Induced Responses of Plants at the Metabolic,
Physiological, and Whole-Plant Levels
to
o
o
oo
H
6
o
o
H
O
O3 Effects: V, Decrease; A, Increase; O, No Significant Effect.
Co2 Effects: A, Additive/Synergistic; T, Antagonistic/Ameliorative; O, No Significant Effect.
Plant response
Growth, Yield (cont'd)
Total Biomass (cont'd)
Seed/grain/fruit/tuber
yield
O3 Response3 CO2 Codification"
0 OD
0 OD
V T
V T
V T
V T
0 0
V T
[V] V
V [T]
Species Facility0
Aleppo pine (Pinus halepensis) CEC, P
Scots pine (P. sylvestris) OTC, G
Soybean (G. max) OTC, P
OTC, G
OTC, G
Bean (P. vulgaris) OTC, P
Tomato (L. esculentum) CSTR, P
Potato (S, tuberosum) OTC, G
OTC, G
Wheat (T. aestivum) OTC, G
OTC, G
CEC,P
OTC, G
OTC, G
OTC, G
Wheat (T. aestivum) (cont.) OTC, G
OTC, G
Reference
Kytoviita et al. (1999)
Broadmeadow and Jackson (2000)
Fiscusetal. (1997);
Mulchi et al. (1992);
Mulchietal. (1995)
Heagle et al. (2002)
ReinertandHo(1995)
Finnan et al. (2002)
Persson et al. (2003)
Bender etal. (1999);
Fangmeier et al. (1996);
McKeeetal. (1997a);
Mulchietal. (1995);
Mulholland etal. (1998a)
Rudorffetal. (1996b)
Fangmeier et al. (1996)
Mulholland et al. (1998b, 1998a)
o
HH
H
W
-------
3
Table 9-12 (cont'd). Effects of Increased Carbon Dioxide on Ozone-Induced Responses of Plants at the Metabolic,
Physiological, and Whole-Plant Levels
to
o
o
VO
1
VO
o
s
H
6
O
0
H
O
O
0
o
H
W
Co2 Effects:
Plant response O3 Response3
Growth, Yield (cont'd)
Relative growth rate V
V
V
V
o
o
o
V
[V]
V
o
Specific leaf area-SLA V
V
V
O3 Effects: V, Decrease; A, Increase; O, No Significant Effect.
A, Additive/Synergistic; T, Antagonistic/Ameliorative; O, No Significant Effect.
CO2 Codification"
V
T
T
T
OD
OD
OD
V
V
V
OD
OD
T
OD
Species
Wheat (T. aestivum)
Agropyron smithii
Koeleria cristata
Bouteloua curtipendula
Schizachyrium scoparium
European ash (F. excelsior)
Trembling aspen
(P. tremuloides)
Red oak (Q. rubra)
Durmast oak (Q. petraea)
Scots pine (P. sylvestris)
Radish (R. sativus)
Soybean (G. max)
Facility0
CEQP
CEQP
CEQP
CEQP
CEQP
CEQP
OTQG
CEQP
CEQP
OTQG
OTQG
CEQP
OTQP
OTQG
Reference
Barnes etal. (1995)
Cardoso-Vilhena and Barnes (2001)
Volin etal. (1998)
Volin etal. (1998)
Volin etal. (1998)
Volin etal. (1998)
Broadmeadow and Jackson (2000)
Volin etal. (1998)
Volin etal. (1998)
Broadmeadow and Jackson (2000)
Broadmeadow and Jackson (2000)
Barnes and Pfirrmann (1992)
Reid etal. (1998)
Mulchi etal. (1992)
-------
Table 9-12 (cont'd). Effects of Increased Carbon Dioxide on Ozone-Induced Responses of Plants at the Metabolic,
Physiological, and Whole-Plant Levels
to
o
o
H
6
o
o
H
O
Plant response
Growth, Yield (cont'd)
Specific leaf area-SLA
(cont'd)
Root/shoot ratio
Co2 Effects:
O3 Response3
V
A
A
A
O
A
A
A
V
V
V
0
A
O3 Effects: V, Decrease; A, Increase; O, No Significant Effect.
A, Additive/Synergistic; T, Antagonistic/Ameliorative; O, No Significant Effect.
CO2 Codification"
A
A
A
A
O
O
A
A
V
A
T
OD
A
Species
White clover (T. repens)
(O3-sensitive)
White clover (T. repens)
(O3-tolerant)
Agropyron smithii
Koeleria cristata
Bouteloua curtipendula
Schizachyrium scoparium
Trembling aspen
(P. tremuloides)
Red oak (Q. rubra)
Radish (R. sativus)
Alfalfa (M. sativa)
White clover (T. repens)
(O3-sensitive)
White clover (T. repens)
(O3-tolerant)
Wheat (T. aestivum)
Facility0
CSTR, P
CSTR, P
CEC,P
CEQP
CEQP
CEQP
CEQP
CEQP
CEQP
CEQP
CSTR, P
CSTR, P
CEQP
Reference
Heagleetal. (1993)
Heagleetal. (1993)
Volinetal. (1998)
Volinetal. (1998)
Volinetal. (1998)
Volinetal. (1998)
Volinetal. (1998)
Volinetal. (1998)
Barnes and Pfirrmann (1992)
Johnson etal. (1996a)
Heagleetal. (1993)
Heagleetal. (1993)
McKeeetal. (1997a)
o
HH
H
W
-------
3
Table 9-12 (cont'd). Effects of Increased Carbon Dioxide on Ozone-Induced Responses of Plants at the Metabolic,
Physiological, and Whole-Plant Levels
to
o
o
VO
>^
^
o
H
6
O
0
H
O
Co2 Effects:
Plant response O3 Response3
Growth, Yield (cont'd)
Root/shoot ratio V
(cont'd)
O
O
0
0
Foliar injury A
A
A
A
A
A
Oo m < 015 nnm
O3 Effects: V, Decrease; A, Increase; O, No Significant Effect.
A, Additive/Synergistic; T, Antagonistic/Ameliorative; O, No Significant Effect.
CO2 Codification"
V
OD
OD
OD
OD
V
V
V
V
V
T
Species
Timothy (P. pratense)
Black cherry (P. serotina)
Green ash (F. pennsylvanica)
Yellow poplar (L. tulipifera)
Aspen (P. tremuloides)
Potato (S. tuberosum)
Bean (P. vulgaris)
Cotton (G. hirsutum)
Wheat (T. aestivum)
Trembling aspen
(P. tremuloides)
European beech (F. sylvatica)
Facility0
CEC,P
CSTR, P
CSTR, P
CSTR, P
OTC, G
OTC, G
OTC,P
OTC,P
CEC,P
OTC, G
FACE, G
CEC,P
Reference
Johnson etal. (1996a)
Loats and Rebbeck (1999)
Loats and Rebbeck (1999)
Loats and Rebbeck (1999)
Dicksonetal. (2001)
Donnelly etal. (200 Ib);
Persson et al. (2003)
Heagle et al. (2002)
Heagle etal. (1999)
Barnes etal. (1995)
Mulholland et al. (1997a)
Karnosky etal. (1999)
Wustman etal. (2001)
Grams etal. (1999)
O
H
W
b CO2-modifications of O3-effects resulting from ~2* present levels. (Trends are shown in brackets. Pronounced changes with ontogeny are, for example,
indicated thus: OD-T.)
0 Exposure facilities used: CEC: controlled environment chambers; CSTR: continuously stirred tank reactors (Heck etal. 1978); FACE: free air CO2
enrichment facilities; OTC: open-top chambers. G: plants rooted in the ground; P: plants grown in pots. All species are C3 except corn, Bouteloua
and Schizachyrium.
-------
1 the evidence in Table 9-12 that elevated CO2 levels can ameliorate the inhibition of growth
2 caused by O3 in many species, although the precise balance among the mechanisms involved
3 may well vary from species to species. Three important caveats must be raised with regard to
4 the findings presented in Table 9-12:
5 • the applicability of results from experiments with an abrupt (step) increase in CO2 level
to understanding the consequences of the gradual increase in CO2 predicted for the
troposphere over the next hundred years;
g • the validity of the findings in several long-term studies (particularly with tree species)
conducted using potted plants, because of possible added stresses imposed on their root
systems relative to trees growing in the field; and
7 • the relevance to understanding the effects of climate change of studies focused solely on
CO2 enrichment at current ambient conditions of temperature and precipitation patterns
that provide no insights into possible interactive effects as these other climatic variables
change concurrently with increasing CO2 (IPCC, 2001).
8 The first caveat concerns the distinctly different natures of the exposures to O3 and CO2
9 experienced by plants in the field. Changes in the ambient concentrations of these gases have
10 very different dynamics. In the context of climate change, CO2 levels increase relatively slowly
11 and may change little over several seasons of growth. On the other hand, O3 presents a
12 fluctuating stress with considerable hour-to-hour and day-to-day variability (Polle and Pell,
13 1999). Almost all of the evidence presented in Table 9-12 comes from experimentation
14 involving plants grown from the outset in, or subjected to, an abrupt or step increase to a higher
15 (more or less double), steady CO2 concentration. In contrast, the O3 exposure concentrations
16 usually varied from day to day. Luo and Reynolds (1999), Hui et al. (2002), and Luo (2001)
17 noted the difficulties in predicting the likely effects of a gradual CO2 increase from experiments
18 involving a step increase or those using a range of CO2 concentrations. Indeed, although using
19 the much accelerated time-scale of an 80-day growing season, Hui et al. (2002) clearly showed
20 significant differences between the rates and magnitudes of various physiological and growth
21 responses of plantain (Plantago lanceolatd) to CO2 between gradual and step increase
22 treatments. The authors concluded that, even though there were major differences in most of the
23 parameters studied between the gradual and step treatments, "the convergence of the measured
24 parameters at the end of the experiment provides some encouragement for the applicability of
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1 step-type experiments in the field; however, the study suggests caution in interpreting early
2 results from short-term studies."
3 In long-term studies, the matter of photosynthetic acclimation to elevated CO2 levels has to
4 be considered. Lawlor and Keys (1993) define acclimation in terms of long-term (days, weeks),
5 irreversible physiological changes, in contrast to regulation, which relates to more rapid
6 (minutes, hours), reversible changes. Each may be positive or negative, but many studies
7 indicate that, while positive acclimation to elevated CO2 levels initially led to enhanced
8 photosynthesis and growth, negative acclimation ultimately ensued and reduced CO2
9 assimilation and growth rates. However, the consensus from recent studies and reviews is that
10 such negative acclimation is most likely to occur in situations in which plants are grown under
11 some additional stress, induced, e.g., by limitations to growth posed by lack of resources such as
12 water or nutrients. The meta-analysis by Curtis (1996) revealed that slow or little negative
13 acclimation was noted in studies on unstressed tree species with unhindered opportunities for
14 root growth and development, a view originally suggested by Arp and Drake (1991) and largely
15 supported in the review by Eamus (1996). A nonwoody perennial, the rhizomatous wetland
16 sedge, Scirpus olneyi, grown in its natural environment with no edaphic limitations showed no
17 negative acclimation after 4 years; in fact, photosynthetic capacity increased by 31% (Arp and
18 Drake, 1991). No negative acclimation of well-watered, field-grown Ponderosa pine (Pinus
19 ponderosd) trees was observed by Tissue et al. (1999) after 6 years of growth at 2x-ambient CO2
20 levels. Gifford and Morison (1993) have summarized the situation thus: "Where the aerial or
21 root environment for a plant is restricted (as with inter-plant competition, for example), positive
22 feedback is limited and adjustments to the changed resource input balance under high CO2 can
23 include 'down-regulation' of leaf photosynthesis rate as an integral part of a positive growth
24 response."
25 The influence of other environmental stress factors is borne out by several long-term tree
26 studies. After 3 years in 565 ppm CO2 in the Duke Forest free-air CO2 enrichment (FACE)
27 facility in North Carolina, maturing loblolly pine (Pinus taedd) trees showed only a marginal
28 CO2-induced carbon gain if grown on a nutritionally moderate site, but zero gain if grown on a
29 nutritionally poor site (Oren et al., 2001). This is in sharp contrast to the substantially increased
30 initial growth rates in elevated CO2 reported by DeLucia et al. (1999), but it is supported by the
31 observations of Tognetti et al. (2000) on five Mediterranean tree species growing for many years
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1 adjacent to geothermal springs releasing CO2 to provide ambient levels averaging 700 ppm.
2 No significant differences in radial growth of the oaks (Quercus cerris, Q. ilex, and
3 Q. pubescens), strawberry tree (Arbutus Unedo), and flowering ash (Fraxinus Ornus) could be
4 detected between trees at the naturally enriched site and those at a nearby site exposed to normal
5 ambient CO2 (-350 ppm). The authors concluded that limited availability of water and nutrients
6 may have counteracted any positive effects of CO2 on growth at the enriched site or that the trees
7 had acclimated to the higher CO2 levels.
8 Because the ameliorative effects of CO2 on responses to O3 (Table 9-14) were reported
9 mostly in short-term studies involving an abrupt increase in CO2 level, it is appropriate to ask
10 whether this amelioration is likely to persist to a time when the ambient CO2 concentration is
11 relatively stable at such levels. Regardless of any negative acclimation due to resource
12 limitations that may occur in the interim, steadily rising CO2 levels may well lead to natural
13 selection and genetic change. Nevertheless, it seems reasonable to expect that the amelioration
14 of O3 impact at elevated CO2 levels will be maintained in many situations, but the negative
15 acclimation that will probably occur in situations where other resources become limiting will
16 reduce the degree of protection.
17 Another caveat regarding the validity of some of the observations in Table 9-12 is related
18 to the matter of stress-induced negative acclimation to elevated CO2 and concerns results
19 obtained using potted plants. Although much of the recent information on CO2 effects has come
20 from experiments with plants rooted in the ground, more than half of the studies listed in
21 Table 9-12 used potted plants, whether in controlled environment and greenhouse chambers or in
22 OTCs. The recent meta-analysis of data on the effects of elevated CO2 on soybean (Glycine
23 max) physiology and growth by Ainsworth et al. (2002) revealed a threefold smaller stimulation
24 of seed yield in pot-grown than in field-grown plants, even when large (> 9 L) pots were used.
25 Loats and Rebbeck (1999) noted that their inability to observe any CO2-induced increases in the
26 root/shoot ratios in seedlings of three broad-leaved tree species (Table 9-13) may have resulted
27 from their use of potted plants. The use of potted plants was a confounding factor in the studies
28 of Taylor et al. (2001) of the differences in leaf growth of poplar (Populus) hybrids between
29 plants exposed to elevated CO2 in controlled environment chambers (potted plants), OTCs or a
30 FACE facility. Eamus (1996) has suggested that any long-term observations of reduced stomatal
31 conductance are almost universally a consequence of pot-based trials at elevated CO2 with
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1 restricted root growth, in contrast to the lack of such a decline observed with trees rooted in the
2 ground. Although the majority of the cases cited in Table 9-12 indicate that O3 and CO2 act
3 additively or synergistically in causing stomatal closure, there are numerous exceptions.
4 Any reduction in stomatal aperture has consequences other than merely restricting O3
5 uptake and the exchange of other gases. In particular, the rate of transpiration is reduced and,
6 while this tends to increase water-use efficiency, it may also lead to decreased mineral uptake,
7 which could adversely impact growth over extended periods. Furthermore, less transpiration
8 also means less evaporative cooling and an increase in leaf temperature, independent of any
9 change in mean air temperatures.
10 Hence, the final caveat regarding Table 9-12 concerns the interactions of O3 and CO2 with
11 other climatic variables, especially mean temperature. In light of the key role played by
12 temperature in regulating physiological processes and modifying plant response to increased
13 CO2 levels (Long, 1991; Morison and Lawlor, 1999) and the knowledge that relatively modest
14 increases in temperature may lead to dramatic consequences in terms of plant development
15 (Lawlor, 1998), it is unfortunate that much of the large investment in time and resources spent
16 on recent studies of the effects of climate change on vegetation have gone into investigations
17 limited to increasing our knowledge of the effects of higher levels of CO2 at current ambient
18 temperatures.
19 Some attention is now being paid to investigating the concurrent effects of CO2 increases
20 and warming (recently reviewed by Rowland-Bamford, 2000 and Morison and Lawlor, 1999),
21 but the observed interactive effects on plant growth are inconsistent. For example, a FACE
22 study with ryegrass (Lolium perenne) ahowed that increased temperatures (provided by infrared
23 heaters) reduced the dry matter gain resulting from increased CO2 levels (Nijs et al., 1996). The
24 field studies by Shaw et al. (2002) on a California annual grassland dominated by the grasses
25 Avena barbata and Bromus hordeaceus and the forbs Geranium dissectum and Erodium botrys
26 involved free-air increased CO2 as well as increased temperature, precipitation, and N supply.
27 Not only did increased temperature reduce CO2-stimulated net primary productivity (NPP), but
28 increased CO2 itself, combined with other factors, was found to be able to cause reduced NPP.
29 There have been several investigations of effects on wheat (Triticum aestivum). Batts et al.
30 (1997) used plastic tunnels to create temperature gradients and maintain elevated CO2 levels
31 over field-grown wheat and found that, in each of 4 years of study, a temperature rise of-1.5 °C
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1 consistently canceled the growth and yield increases caused by a doubling of the CO2 level
2 above ambient. Similar findings were reported by van Oijen et al. (1999) and van Oijen and
3 Ewart (1999) in OTC field studies. Half of the chambers were cooled 1.6 to 2.4 °C below the
4 uncooled chambers, to cancel out the normal temperature increase over ambient, due to the
5 so-called "chamber effect" (usually a 1 to 3 °C increase above ambient temperature (Heagle
6 et al., 1988). Although temperature had no effect on CO2-enhanced assimilation rates, the
7 CO2-enhanced growth and grain yields observed in the cooled chambers were effectively
8 canceled out in the warmer chambers. The authors attributed this effect to accelerated
9 phenology, a shorter period for grain filling, and a lower leaf area index (LAI; total leaf area per
10 unit ground area) in the warmer chambers. Wheeler et al. (1996) observed that the benefit to
11 wheat of doubling the CO2 level was offset by a mean seasonal increase of only 1 to 1.8 °C.
12 With the continuing use of OTCs for field research, Runeckles (2002), has suggested that the
13 temperature rise due to the chamber effect in OTCs should be exploited (and measured) as a
14 means of exploring temperature x CO2 as well as temperature * CO2 * O3 interactions.
15 An indirect affirmation of the importance of temperature as a component of climate change
16 on wheat yield was provided by van Oijen and Ewart (1999), using two simulation models,
17 AFRCWHEAT2-O3 and LINTULCC (Ewart et al., 1999). They analyzed data from the
18 ESP ACE-wheat program, which involved 25 OTCs experiments in 1994, 1995, and 1996 at nine
19 European locations (Jager et al., 1999). Both models were able to predict control-treatment grain
20 yields closely (5.5 ±1.2 and 5.8 ±1.2 t.ha"1, respectively, versus the observed 5.9 ±1.9 t.ha"1),
21 and both indicated a predominantly negative effect of temperature on the yield response to
22 increased CO2 (a 3 °C rise reduced the gain in yield from 30 to 14%). However, neither model
23 had an R2 > 0.3, indicating that the models included other sources of variability among the sites
24 than the climatic factors. The multiple linear regression developed by Bender et al. (1999) based
25 on the same data sets also included temperature as a highly significant covariant. Both studies
26 are discussed more fully below.
27 Other studies, however, have found positive temperature-related growth effects, as
28 suggested by the early Idso and Idso (1994) analysis. In an OTC study using the perennial grass
29 Festucapratensis in which a temperature increase of 3 °C above ambient was combined with
30 CO2-enrichment to 700 ppm, both CO2 and temperature caused increases in total above-ground
31 biomass (Hakala and Mela, 1996). Studies with potato (Cao et al., 1994) and soybean (Ziska and
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1 Bunce, 1997) using potted plants in controlled environment chambers also showed temperature-
2 enhanced increases in growth in enriched CO2 atmospheres. Read and Morgan (1996) compared
3 the effects of enriched CO2 and temperature on two grasses: cool-season Pascopyrum smithii
4 and warm-season Bouteloua gracilis. In the latter (a C4 species), 750 ppm CO2 resulted in
5 increased dry matter production at daytime temperatures as high as 35 °C, but in P. smithii
6 (a C3 species), CO2-stimulated growth was greatest at 20 °C. However, the stimulation was
7 progressively attenuated by increased temperature, so that at 35 °C, growth in 750 ppm was only
8 one third of that in 350 ppm CO2 at 20 °C.
9 Although the picture we have of temperature x CO2 interactions is inconsistent, Rowland-
10 Bamford (2000) has provided persuasive evidence that the nature of response to temperature in
11 the grain yield of crops with as different temperature optima as rice and wheat will depend upon
12 whether the change is above or below the temperature optimum.
13 But what if we add O3 as another variable? Unfortunately, there have been very few
14 studies of the three-way interaction. With the information available on CO2 x O3 interactions
15 (Table 9-12) and the limited information on temperature x O3 interactions (discussed in Section
16 9.4.4.2) simulation modeling can attempt to provide estimates of O3 x CO2 x temperature effects,
17 but experimental observation is still required to validate the models. The questions that need to
18 be answered are: if increased temperature can offset the gains in productivity afforded by
19 increased CO2 in important species such as wheat, and increased CO2 can offset the reductions in
20 productivity caused by O3, will increased temperature modify this protective effect. And if so, in
21 what manner.
22 To date, the only information available appears to consist of the reports by van Oijen and
23 Ewart (1999) and Bender et al. (1999) referred to above. In the former's simulation studies, the
24 overall yield depression of wheat caused by O3 was found to be 7 ± 4% for both
25 AFRCWHEAT2-O3 and LINTULCC models versus an observed 9 ± 11%. The enhancements
26 due to CO2 were predicted to be 24 ± 9% and 42 ± 11%, respectively, which straddled the
27 observed 30 ± 22% gain. Based on the 13 experiments that included all four treatments (±O3,
28 ±CO2), an actual 10% yield loss due to O3 at ambient CO2 levels was reduced to a 4% loss by the
29 elevated CO2. The AFRCWHEAT2-O3 model predicted 7 and 4% losses, and LINTULCC
30 model predicted 8 and 5% losses due to the O3 and O3 + CO2 treatments. The actual and
31 simulated yield increases in response to CO2 increased further with increasing temperature, but
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1 although temperature had no discernible effect on the observed depression of yield caused by O3
2 alone, both models suggested that the yield reduction was diminished both by higher
3 temperatures and higher CO2 levels.
4 The analysis of the ESPACE-wheat experiments by Bender et al. (1999) led to the
5 following multiple linear regression:
6
7
8 7=1004.6*** + 0.588***[C02]- 1.908**[O3]- 31.230***[r]+ 7.309[/]- 1448.423***[H2O], (9-5)
9
10
11 where 7= grain yield, g-m~2; [CO2] = ppm CO2; [O3] = ppb O3, 12-h mean; [7] = °C; [7] = light
12 intensity, MJ-nT2/day; and [H2O] is a dummy variable: well watered = 1; limited water
13 supply = 2. (***, p < 0.001; **, p < 0.01; the coefficient for/was not significant.) With
14 R2 = 0.3983, adjusted for 258 degrees of freedom, a large part of the variability was still
15 unaccounted for by the five variables. However, this analysis suggests that CO2, O3,
16 temperature, and water-status are important codeterminants of wheat yield but assumes no
17 interactions. Substitution in the model at summer light intensities and with well-watered plants
18 indicates that, at 20 °C, a doubling of CO2 levels to 700 ppm alone, would lead to a 29.5%
19 increase in yield, while 50 ppb O3 alone would decrease yield by 10.9%. With both gases at
20 those levels, the yield would only increase 20%, but with a concurrent temperature rise of 2 °C,
21 it would shrink to a 9.6% increase.
22 Both studies, therefore, indicate an amelioration of the effects of O3 by CO2, the magnitude
23 of which would be reduced at warmer temperatures. However, they relate to a single crop whose
24 response to CO2 is temperature-sensitive. Information about other species in which the effects of
25 CO2 and temperature are additive are limited. However, Wolf and Van Oijen (2003) recently
26 described a model (LPOTCO) simulating the effects of changes in climatic variables, CO2 and
27 O3 on tuber yield potential of irrigated potato (Solanum tuberosum cv. Bintje) over locations
28 within the European Union ranging from Finland to Italy. They noted that although increased
29 CO2, O3, and light intensity were predominant controlling factors, increased temperature also
30 influenced potential yields substantially, with increases in northern latitudes (attributed to a
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1 longer growing season) but decreases in southern latitudes (attributed to decreased assimilate
2 production).
3 A clear understanding of the complex interactions of increased CO2 and temperature with
4 O3 must await further experimentation or simulations. However, it seems likely that any CO2-
5 induced amelioration of the adverse effects of O3 on aspects of growth other than seed or grain
6 yield may be lessened or increased by increased temperature, depending upon the temperature,
7 optima for the species, along the lines suggested by Rowland-Bamford (2000).
8 Other crop simulation models which incorporate O3 and some of the various environmental
9 factors, including elements of climate change, have been reviewed by Kickert et al. (1999) and
10 Rotter and van de Geijn (1999). However, to date, the applications tend to have focused on
11 interactions of O3 with factors such as soil moisture or nutrient availability.
12 With forest trees, the situation has the added complexity of a perennial growth form and
13 the inevitability, over time, of subjection to additional environmental stresses such as nutrient-
14 limitation. Here, too, although numerous models of tree growth have been described, there
15 appear to have been few applications to interactions of O3 and factors of climate change.
16 Constable et al. (1996) used TREEGRO to model the growth of Ponderosa pine (Pinus
17 ponderosd) exposed to three O3 levels (0.5x, l.Qx, and 2x ambient), two levels of CO2 (ambient
18 and ambient + 200 ppm CO2), and two temperature regimes (ambient and ambient + 4 °C). Plant
19 growth was predicted to be decreased 1, 19, and 39% by the three levels of O3, respectively.
20 Increased CO2 reduced the loss at the highest O3 level to 7% but the combination of elevated
21 CO2 with the higher temperature more than overcame the adverse effects of O3, leading to a 4%
22 increase, largely attributed to increased fine root mass. The authors suggested that, in relation to
23 the baseline conditions used in the simulations (Corvallis, OR), higher concentrations of CO2 and
24 O3 and a warmer climate will have little impact on total-tree growth, but they noted the
25 importance of undertaking multiple stress studies in order to be able to make accurate forecasts
26 of the impact of such changes on forest trees.
27 More recently, Constable and Friend (2000) compared the capabilities of six published
28 process-based models (CARBON, ECOPHYS, PGSM, TRE-BGC, TREEGRO, and W91) for
29 simulating tree response to elevated CO2, O3, and temperature. They concluded that although
30 these models were capable integrators of the effects of various environmental factors on
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1 individual processes such as photosynthesis, they were less reliable when extrapolating to
2 growth.
3 Although most of the research emphasis has been on simple CO2 x O3 interactions, a few
4 isolated studies of interactions have involved O3, CO2, and biotic environmental factors. Heagle
5 et al. (1994) observed that both O3 and CO2 tended to be additive in encouraging the growth of
6 spider mite (Tetranychus urticae) populations on clover (Trifolium repens). Infection of wheat
7 (Triticum aestivum) with leaf rust (Puccinia reconditd) sensitized the plants to O3 injury, but its
8 severity was significantly reduced in elevated CO2 (Tiedemann and Firsching, 2000). The
9 effects of O3 and CO2 on mycorrhizal symbioses was studied by Kytoviita et al. (1999) who
10 found that CO2 did not ameliorate the adverse effects of O3 on the root growth of Aleppo pine
11 (Pinus halepensis) and birch (Betulapenduld). In another study with Aleppo pine, Kytoviita
12 et al. (2001) noted that both O3 and elevated CO2 reduced mycorrhiza-induced N-uptake by the
13 roots. In Scots pine (Pinus sylvestris\ Kasurinen et al. (1999) observed transient effects of
14 elevated CO2 and O3 on root symbiosis, but none of the effects persisted over the 3 years of
15 the study.
16 The soil water x O3 x CO2 interaction was experimentally investigated by Broadmeadow
17 and Jackson (2000) in Durmast oak (Quercuspetraea), European ash (Fraxinus excelsior), and
18 Scots pine (Pinus sylvestris). No interactions were noted with ash and pine; but with oak,
19 elevated CO2 ameliorated and irrigation exacerbated the effects of O3, although the resultant
20 effects were essentially additive.
21 Booker (2000) noted that soil nitrogen levels interacted only slightly with O3 and CO2 in
22 determining the composition of cotton (Gossypium hirsutum) leaves and roots. Carbon dioxide
23 reversed the inhibition of leaf growth caused by O3, but increased N-fertility tended to reduce
24 this reversal.
25 Because of the small number of studies of possibly significant interactions of three or more
26 environmental factors, it impossible to draw any sweeping conclusions as to how O3, in the
27 context of global climate change, may affect relationships among plants and insects, diseases,
28 and symbionts or among plants and nutrients or other air pollutants. The only interaction that
29 has some degree of general support is the amelioration of adverse O3 effects by elevated CO2.
30
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1 9.4.8.2 Ozone-UV-B Interactions
2 As noted in the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996), depletion
3 of stratospheric O3 by halofluorocarbons has resulted in increased intensities of ultraviolet-B
4 (UV-B) radiation (280 to 320 nm wavelengths) at the earth's surface. The situation is discussed
5 more fully in Chapter 10.
6 While stratospheric O3 depletion may result in increased surface UV-B irradiation,
7 absorption of UV-B is a property of the O3 molecule regardless of its location; and surface UV-B
8 flux is, therefore, also reduced by O3 in the troposphere. Although only about 10% of the total
9 atmospheric O3 column occurs in the troposphere (Fishman et al., 1990), it contributes a
10 disproportionately greater absorption effect than stratospheric O3 because the UV radiation
11 penetrating the troposphere becomes increasingly diffuse as it reaches the surface, with a
12 consequent increase in mean path length (Briihl and Crutzen, 1989). Any benefits to vegetation
13 from reduced ambient O3 stress must, therefore, also be viewed in the context of possible
14 adverse effects due to increased UV-B irradiation. There are, thus, two distinct types of possible
15 interactions between surface level O3 and UV-B radiation:
16 • direct interactions involving simultaneous, sequential, or mixed exposures to O3 and
UV-B stresses; and
17 • effects on responses to UV-B itself resulting from changes in radiation intensity caused
by changes in surface level O3 concentrations.
18 Only the first type of interaction is discussed here. The second type of interaction has broad
19 implications for both health and welfare and focuses on the impacts of UV-V radiation per se.
20 It is dealt with separately in Chapter 10, which includes a critical review of the experimental
21 difficulties faced in undertaking meaningful plant research in order to reach a clear
22 understanding of the effects of increased UV-B radiation, the evidence for both its adverse and
23 beneficial effects on plants, and the potential for changes in ambient surface O3 levels to modify
24 these effects.
25 The most recent reviews specifically addressing the combined effects of tropospheric O3
26 and UV-B on plants are by Runeckles and Krupa (1994) and Krupa and Jager (1996), although
27 the topic has also been included in several more general reviews of O3 effects and factors of
28 climate change such as those by Unsworth and Hogsett (1996), Krupa et al. (1998), Posthumus
29 (1998), Groth and Krupa (2000), and Krupa and Groth (2000).
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1 However, little new information has become available since Runeckles and Krupa (1994)
2 noted that the scanty knowledge of the effects of UV-B and O3 combinations available at that
3 time was derived solely from studies of soybean (Glycine max). Miller et al. (1994) observed no
4 interaction and no effect of UV-B on yield, in contrast to a previous report by Teramura et al.
5 (1990) using the same cultivar, Essex. More recently, in a study of the saltmarsh grass Elymus
6 athericus subjected to reciprocal exposures to O3 and UV-B, van de Staaij et al. (1997) observed
7 no interactive effects and no adverse effects of UV-B following 14-day exposures, even though
8 an earlier report showed that longer exposures to UV-B (65 days) could cause a 35% loss of
9 biomass (Van De Staaij et al., 1993). However, in a study in which ambient, high altitude UV-B
10 levels were compared with near zero levels, at ambient or 2x-ambient levels of O3, interactions
11 involving the levels of antioxidants in Norway spruce (Picea abies) and Scots pine (Pinus
12 sylvestris) were reported by Baumbusch et al. (1998). Schnitzler et al. (1999) subsequently
13 reported that O3-induced injury and adverse effects on photosynthesis were more pronounced
14 with near zero UV-B levels, indicating an amelioration of the O3-response. A later study with
15 Scots pine (Zinser et al., 2000) revealed O3 * UV-B interactions at the gene expression and
16 biochemical levels. In contrast, Ormrod et al. (1995) reported that UV-B predisposed
17 Arabidopsis thalliana to injurious growth effects of O3.
18 At various organizational levels, Runeckles and Krupa (1994) identified several similarities
19 between plant response to O3 and UV-B, and at the level of gene expression there have recently
20 been several reports of both similarities and distinctions. Willekens et al. (1994) reported similar
21 effects of O3, UV-B, and SO2 on the expression of antioxidant genes in Nicotiana
22 plumbaginifolia. In parsley (Petroselinum crispum\ Eckey-Kaltenbach et al. (1994a) found that
23 O3 was a cross-inducer for both the UV-B-induced enhanced biosynthesis of flavonoids and the
24 pathogen-induced furanocoumarin phytoalexins, in keeping with the previously observed
25 O3-induction of fungal and viral defense reactions. In this regard, Yalpani et al. (1994) provided
26 evidence that, in tobacco (Nicotiana tabacum), O3 and UV-B acted similarly in increasing
27 disease-resistance via a salicylate-mediated enhancement of defense proteins. However,
28 subsequent work with tobacco led Thalmair et al. (1996) to conclude that exposure to UV-B did
29 not lead to the accumulation of pathogenesis-related proteins. In Scots pine (Pinus sylvestris),
30 although O3 is known to induce stilbene synthase and cinnamyl alcohol dehydrogenase, UV-B
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1 was found to enhance the former but suppress the latter, revealing an interaction at the level of
2 gene expression (Zinser et al., 2000).
3 In summary, the present base of information about possible interactions between increased
4 UV-B radiation and O3 is insufficient to draw any firm conclusions in terms of gross effects, but
5 there is some evidence of similarities in the effects of O3 and UV-B individually and of the
6 mechanisms involved at the molecular level.
7
8 9.4.8.3 Interactions of Ozone with Multiple Climate Change Factors
9 Despite of the need for experimental investigations of three-way or more complex
10 interactions among O3, CO2, UV-B, temperature, and other climate change factors, few studies
11 have been reported, even without O3 as a factor. In an isolated report, using tomato
12 (Lycopersicon esculentuni) seedlings, Hao et al. (2000) employed preexposure to UV-B (±CO2
13 enrichment) followed by exposure to O3 (± CO2 enrichment). They observed that CO2 more than
14 overcame the inhibition of photosynthesis caused by O3, but pretreatment with UV-B reduced
15 the resultant increase.
16 In view of the unexpected observations made in their grassland study of the combined
17 effects of CO2, temperature, precipitation, and N-supply, Shaw et al. (2002) affirmed that:
18 "Ecosystem responses to realistic combinations of global changes are not necessarily simple
19 combinations of the individual factors." The addition of O3 to the list of variables results in
20 further complexity.
21 Although computer simulation modeling may ultimately lead to improved understanding of
22 these complex issues, to date, no such models appear to have been applied to these interactions,
23 possibly because of the scarcity of experimental data for parameterization.
24
25 9.4.9 Summary - Environmental Factors
26 Although O3 and other photochemical oxidants are phytotoxic, their actions on vegetation
27 may be modified by a host of biotic and abiotic factors in the environment; conversely, they may
28 modify plant response to these other factors. The extensive review of these biological, physical,
29 and chemical factors conducted for the 1996 O3 AQCD (U.S. Environmental Protection Agency,
30 1996) concluded with a statement that our understanding was too fragmented to permit drawing
31 many general conclusions. With today's increased awareness of the need for more complete
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1 information on interactions, it is unfortunate that, in the interval since the 1996 criteria
2 document, rigorous, systematic investigations of interactions have been rare, and most of the
3 new information is as fragmented as before. This is inevitable partly in view of the vast scope of
4 the possible interactions between O3 and other environmental variables. But it is also an
5 unavailable outcome of having the personal interests (and funding) of individual groups
6 determine the studies to be undertaken, instead of a coordinated program of research focused on
7 systematic investigations to improve our understanding our ability to assess the risks posed by
8 photochemical oxidants to cultivated and natural vegetation.
9 In the area of biotic interactions, new evidence with regard to insect pests and diseases has
10 done little to remove the uncertainties noted in the 1996 criteria document. Most of the large
11 number of such interactions that may affect crops, forest trees, and other natural vegetation have
12 yet to be studied. The trend suggested previously that O3 increases the likelihood and success of
13 insect attack has received some support from recent studies, but only with respect to chewing
14 insects. With the economically important group of sucking insects such as the aphids, no clear
15 trends have been revealed by the latest studies. Hence, although it seems likely that some insect
16 problems could increase as a result of increased O3 levels, we are still far from being able to
17 predict the nature of any particular O3 plant insect interaction, its likelihood, or its severity.
18 The situation is a little clearer with respect to interactions involving facultative,
19 necrotrophic plant pathogens, with O3 generally leading to increased disease. With obligate,
20 biotrophic fungal, bacterial, and nematode diseases, there are twice as many reports indicating
21 O3-induced inhibitions than enhancements. The frequent reports that infection by obligate
22 biotrophs reduces the severity of O3-induced foliar injury should not be interpreted as
23 "protection" because of the negative effects on the host plant of the disease per se. With obligate
24 biotrophs, the nature of any interaction with O3 is probably dictated by the unique, highly
25 specific biochemical relationships between pathogen and host plant. At this time, therefore,
26 although some diseases may become more widespread or severe as a result of exposure to O3, it
27 is still not possible to predict which diseases are likely to present the greatest risks to crops and
28 forests.
29 Several studies have indicated that the functioning of tree root symbioses with mycorrhizae
30 may be adversely affected by O3, but there is also evidence that the presence of mycorrhizae may
31 overcome root diseases stimulated by O3 and that O3 may encourage the spread of mycorrhizae
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1 to the roots of uninfected trees. The latest studies, therefore, present no clearer picture of the
2 likely nature of simple interactions of O3 and root symbionts, but in view of the importance of
3 mycorrhizae as below-ground components of ecosystems, they are discussed more fully in
4 Section 9.5.
5 The few recent studies of the impact of O3 on intraspecific plant competition have again
6 confirmed that grasses frequently show greater resilience than other types of plants. In grass-
7 legume pastures, the leguminous species suffer greater growth inhibition. And the suppression
8 of Ponderosa pine seedling growth by blue wild-rye grass was markedly increased by O3.
9 However, we are far from being able to predict the outcome of the impact of O3 on specific
10 competitive situations, such as successional plant communities or crop-weed interactions.
11 Light, a component of the plant's physical environment, is an essential "resource" whose
12 energy content drives photosynthesis and CO2 assimilation. It has been suggested that increased
13 light intensity may increase the sensitivity to O3 of light-tolerant species while decreasing that of
14 shade-tolerant species, but this appears to be an over-simplification with many exceptions.
15 Temperature affects the rates of all physiological processes based on enzyme-catalysis and
16 diffusion, and each process and overall growth (the integral of all processes) has a distinct
17 optimal temperature range. Although some recent field studies have indicated that O3 impact
18 significantly increases with increased ambient temperature, other studies have revealed little
19 effect of temperature. But temperature is unquestionably an important variable affecting plant
20 response to O3 in the presence of the elevated CO2 levels contributing to global climate change
21 (see below). In contrast, evidence continues to accumulate to indicate that exposure to O3
22 sensitizes plants to low temperature stress by reducing below-ground carbohydrate reserves,
23 possibly leading to responses in perennial species ranging from rapid demise to impaired growth
24 in subsequent seasons.
25 Although the RH of the ambient air has generally been found to increase the adverse
26 effects of O3 by increasing stomatal conductance and thereby increasing O3 flux, abundant
27 evidence indicates that the ready availability of soil moisture results in greater sensitivity to O3.
28 The partial "protection" against the adverse effects of O3 afforded by drought has been observed
29 in field experiments and modeled in computer simulations. There is also compelling evidence
30 that O3 can predispose plants to drought stress. Hence, the response will depend to some extent
31 upon the sequence in which the stresses occur, but, even though the nature of the responses is
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1 largely species-specific, successful applications of model simulations led to larger-scale
2 predictions of the consequences of O3 x drought interactions. However, it must be recognized
3 that regardless of the interaction, the net result on growth in the short-term is negative, although
4 in the case of tree species, other responses such as increased water use efficiency could be a
5 benefit to long-term survival.
6 Mineral nutrients in the soil, other gaseous air pollutants, and agricultural chemicals
7 constitute chemical factors in the environment. The evidence regarding interactions with
8 specific nutrients is still contradictory. Some experimental evidence indicates that low general
9 fertility increases sensitivity to O3, while simulation modeling of trees suggests that nutrient
10 deficiency and O3 act less than additively, but there are too many example of contrary trends to
11 permit any sweeping conclusions. Somewhat analogously with temperature, it appears that any
12 shift away from the nutritional optimum may lead to greater sensitivity, but the shift would have
13 to be substantial before a significant effect on response to O3 was observed.
14 Interactions of O3 with other air pollutants have received relatively little recent attention.
15 The situation with SO2 remains inconsistent, but seems unlikely to pose any additional risk to
16 those related to the individual pollutants. With the NO and NO2, the situation is complicated by
17 their nutritional value as a N source. In leguminous species, it appears that NO2 may reduce the
18 impact of O3 on growth, with the reverse in other species, but the nature of the exposure pattern,
19 i.e., sequential or concurrent, also determines the outcome. Much more investigation is needed
20 before we will be able to predict the outcomes of different O3-NO-NO2 scenarios. The latest
21 research into O3 x acid rain interactions has confirmed that, at realistic acidities, significant
22 interactions are unlikely. A continuing lack of information precludes offering any
23 generalizations about interactive effects of O3 with NH3, HF, or heavy metals. More evidence
24 has been reported that the application of fungicides affords some protective effects against O3.
25 Over the last decade, considerable emphasis has been placed on research into O3
26 interactions with the components of global climate change: increased atmospheric CO2,
27 increased mean global temperatures, and increased surface level UV-B radiation. However,
28 most of these studies have tended to regard increased CO2 levels and increased mean
29 temperatures as unrelated phenomena. Experiments into the effects of doubled CO2 levels at
30 today's mean ambient temperatures are not paricularly helpful in trying to assess the impact of
31 climate change on responses to O3. To date, the limited experimental evidence and evidence
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1 obtained by computer simulation suggest that in a 600+ ppm world, although the enriched CO2
2 would more than offset the impact of O3 on responses as varied as wheat yield or the growth of
3 young Ponderosa pine trees, the concurrent increase in temperature would reduce but probably
4 not eliminate the net gain. A similar decrease in the net gain resulting from the complete
5 reversal by CO2 of the inhibition of photosynthesis caused by O3 has been reported for increased
6 UV-B irradiation. However, these are preliminary results based on minimal data.
7 In conclusion, although the increased use of computer simulations may be important in
8 suggesting outcomes of the many complex interactions of O3 and various combinations of
9 environmental factors, the results obtained will only be as reliable as the input data used for their
10 parameterization. The data needed for good simulations can only come from organized,
11 systematic study. For predicting the future, ignorance is as good as dependence on poor
12 simulations.
13
14
15 9.5 EFFECTS-BASED AIR QUALITY EXPOSURE-AND
16 DOSE-RESPONSE INDICES
17 9.5.1 Introduction
18 An index is needed that relates measured plant damage (i.e., growth) to ambient ozone
19 concentration over time. The index can provide both a consistency for reviewing exposure-
20 response effects in research, as well as a metric for developing a biologically-relevant air quality
21 standard that protects ecological resources. The quantifying function over some time frame has
22 frequently been referred to as "dose-response" and "exposure-response". The distinction being
23 where the pollutant concentration is measured: "Dose" is the measure of the pollutant
24 concentration absorbed by the leaf over some time period, whereas "exposure" is the ambient air
25 concentration measured nearby the plant over some time period.
26 Plant ozone uptake from the ambient air (either rate or uptake or cumulative seasonal
27 uptake) is the ideal measure, because without ozone or its reactive product(s) reaching the target
28 tissue there is no effect. Uptake is controlled in part by stomata (see Section 9.3 for detailed
29 discussion). An uptake measure should integrate all those environmental factors that influence
30 stomatal conductance, e.g., temperature, humidity, soil water status. However, a direct measure
31 of the internal leaf concentration of ozone is technically difficult and thus uptake values are
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1 generally obtained with models that require species- and site-specific variables. Because of this,
2 a surrogate for uptake (i.e., exposure index) was sought early on using statistical summaries of
3 ambient pollutant concentration over some integral of time (O'Gara, 1922; Lefohn and Benedict,
4 1982; Lee et al., 1988; U.S. Environmental Protection Agency, 1986, 1992, 1996).
5 An index of exposure that is biologically relevant must then consider those factors known
6 to modify the plant response by altering ozone uptake (Hogsett et al., 1988; U.S. Environmental
7 Protection Agency, 1996), including the temporal dynamics of exposure (e.g., concentration,
8 frequency, duration), plant phenology (see Section 9.4), plant defense mechanisms (e.g.,
9 antioxidants) (see Section 9.3), and site climate and soil factors (e.g., temperature, vpd, soil
10 moisture) (see Section 9.4). In using these indices to develop air quality standards, the needs of
11 policy makers must also be considered, and those include simplicity and understandability
12 (Fairley and Blanchard, 1991). The development of such indices continues to be a challenge.
13
14 9.5.2 Summary of Conclusions from the Previous Criteria Document
15 The 1996 AQCD (U.S. Environmental Protection Agency, 1996a) focused primarily on
16 development of exposure indices, not flux, to quantify growth and yield effects in crops,
17 perennials and trees (primarily seedlings), and not foliar injury. The testing of the adequacy of
18 these indices to order the measured responses of growth and/or yield in crops and tree species, as
19 seedlings, was accomplished through regression analyses of earlier exposure studies. No direct
20 experimental testing of the range of indices has yet been accomplished. It was recognized that
21 these indices were only surrogates for O3 uptake or dose. Their development focused on
22 consideration and inclusion of some, but not all, factors that affect O3 uptake and expression of
23 effects (e.g., Lee et al., 1988). The 1996 document (U.S. Environmental Protection Agency,
24 1996a) drew a number of conclusions that built on even earlier conclusions (U.S. Environmental
25 Protection Agency, 1986, 1992). These conclusions are still valid today based on a review of
26 research published since 1996.
27 Studies prior to 1996, and after, indicate that the components of exposure, including
28 concentrations, temporal dynamics (e.g., time of day of peak events), frequency of occurrence,
29 duration, and respite time, are integral to developing indices of exposure that relate to growth
30 response. Evidence from the few direct experimental studies of varying exposure components
31 indicate the importance of peak concentrations, occurrence, respite time and the importance of
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1 cumulative the concentrations (Musselman et al., 1983, 1986, 1994; Hogsett et al., 1885; U.S.
2 Environmental Protection Agency, 1996a).
3 Exposure duration influences the degree of plant response. Single season, year-long or
4 multi-year experimental results indicate that greater yield losses occurred when plants were
5 exposed for the longer duration, and that a cumulative-type index was able to better describe the
6 relationship between exposure and yield. Those indices not considering duration, e.g., 7-h
7 seasonal mean concentration index, single peak event index, or the index that cumulates all
8 concentrations (i.e., SUMOO), were unable to adequately describe the relationship. These single
9 event or mean-type indices do not consider the role of duration of exposure and focus either only
10 on the peak event or put too much focus on the lower hourly average concentrations (U.S.
11 Environmental Protection Agency, 1996a).
12 Higher hourly averaged concentrations had a greater affect on plant response. It was
13 concluded that cumulative indices that gave greater weight to higher concentrations relate well
14 with plant response (crops and tree seedlings) and order the treatment means in monotonically
15 decreasing fashion with increasing exposure, based on studies that apply two or more types of
16 exposure regimes with replicate studies of the same species. These indices include, among
17 others: SUM06, W126, AOT40 (U.S. Environmental Protection Agency, 1996a).
18 No studies before or after 1996, have enabled a discrimination among the various
19 weighted, cumulative indices. Various functional weighting approaches were used, including
20 allometric, sigmoid or threshold weighting, and compared for best statistical fit of the plant
21 growth or yield data; but no one functional weighting was favored. For use as an air quality
22 standard, however, the need for simplicity, understandability, and ease of monitoring, favored
23 the cumulative threshold-weighted index (SUM06) (U.S. Environmental Protection Agency,
24 1996a).
25 Since higher concentrations occur primarily during the daylight hours, those indices that
26 differentially weight higher concentrations give greater weight to daylight hour concentrations.
27 This is important since stomatal conductance is usually greatest during the daylight hours,
28 compared to concentrations at night when conductance is usually minimal. Peak concentrations,
29 however, do not occur throughout the day; thus the timing of the peak concentration and
30 maximum plant uptake is critical in determining plant response. An exposure index that
31 incorporated either the daily or seasonal temporal patterns of higher concentration occurrence
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1 with the temporal pattern of individual species' stomatal conductance was not reported in the
2 prior 1996 review (U.S. Environmental Protection Agency, 1996).
3 The relative importance of cumulative peak concentrations (> 0.10 ppm) versus cumulative
4 mid-range concentrations (0.05-0.099 ppm) was questioned based on ambient field exposures of
5 sensitive species. Although controlled experiments provided important evidence that the higher
6 hourly average concentrations should be given greater weight than the mid-level values in
7 developing indices, there was concern that under ambient conditions in the field the higher
8 concentrations did not occur at the time of maximum plant uptake. This coincidence was
9 considered to be the critical factor in determining peak concentration impacts on plants. Based
10 on the evidence at that time, it was not possible to conclude whether the cumulative effects of
11 mid-range concentrations were of greater importance than those of peak hourly average
12 concentrations in determining plant response (U.S. Environmental Protection Agency, 1996a).
13 No direct experimental studies had addressed this question prior to 1996, nor have any since.
14 The composite exposure-response functions for crops and tree seedlings were derived from
15 single and multi-year exposure studies that used modified or simulated ambient exposure
16 profiles. These profiles were typified by episodic occurrence of a large number of high O3
17 concentrations; and this type of pattern is prevalent in many but not all rural agricultural and
18 some forested areas in the United States. Selecting a concentration value from these crop and
19 seedling response models may result in an over or underestimation of growth effects if applied to
20 regions of the country where a different type of exposure pattern is prevalent (U.S.
21 Environmental Protection Agency, 1996a). A multi-component index was suggested that
22 combined the peak-weighted, cumulative index with the number of occurrences of hourly
23 averaged concentrations > 0.10 ppm that might reduce the uncertainty associated with selecting
24 the exposure value for protection based on NCLAN type studies (Lefohn and Foley, 1992;
25 Musselman et al., 1994; U.S. Environmental Protection Agency, 1996a). No direct experimental
26 studies addressed this question prior to 1996, nor have any since.
27 Since 1996, additional research has focused on the time of day when the higher hourly
28 average concentrations occur, the time of day of maximum plant uptake, and the diurnal
29 variability of plant defense mechanisms, and various suggestions as to inclusion of these factors
30 in any one of the peak weighted cumulative exposure indices. A much broader literature has
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1 focused since 1996 on developing indices using flux to relate plant response. These new
2 developments are discussed in the sections that follow.
3
4 9.5.3 Use of Exposure Indices to Establish Exposure-Response Relationships
5 Mathematical approaches for summarizing ambient air quality information in biologically
6 meaningful forms that can serve as surrogates for dose for O3 vegetation effects assessment
7 purposes have been explored for more than 80 years (O'Gara, 1922; U.S. Environmental
8 Protection Agency, 1996). Several of the indices introduced have attempted to incorporate some
9 of the biological, environmental, and exposure factors (directly or indirectly) that influence the
10 magnitude of the biological response and contribute to observed variability (Hogsett et al.,
11 1988). In the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996a), the exposure
12 indices were arranged into 5 categories, including (1) One Event, (2) Mean, (3) Cumulative,
13 (4) Concentration Weighted, and (5) Multi-component and were discussed in detail (Lee et al.,
14 1988). Figure 9-16 illustrates how several of the indices weight concentration and accumulate
15 exposure.
16 The indices were developed with knowledge of O3 modes of action, as well as the role of
17 the exposure components and their adequacy tested using earlier exposure studies, e.g., NCLAN
18 (U.S. Environmental Protection Agency, 1996a; Hogsett et al., 1988). Various components of
19 the exposure-response, including concentration, time of day, respite time, frequency of peak
20 occurrence, plant phenology, predisposition, etc., were weighted with various functions and
21 evaluated on their ability in ordering the regression of exposure vs. growth or yield response.
22 The statistical evaluations for each of these indices were accomplished using growth/yield
23 response data from many earlier exposure studies (e.g., NCLAN). This retrospective approach
24 was necessary because there have been very few studies specifically designed to test the
25 goodness of fit of the indices. The regression approach selects those indices that most properly
26 order and space the treatment means to optimize the fit of a linear or curvi-linear model. This
27 approach provides evidence for the best indices, albeit not as defensible as that from studies with
28 experimental designs and analyses that focus on specific components of exposure.
29 Most of the early retrospective studies reporting regression approaches use data from the
30 NCLAN program or from Corvallis, OR (Lee et al., 1987; Lee et al., 1988; Lefohn and Irving,
31 1988; Tingey et al., 1989) or use data collected in California (Musselman et al., 1988; U.S.
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Q.
Q.
0.15
0.12-
0.09-
o
Q.
X
LU
C
O
o
X
0.06-
0.03-
0.00
8 10
n m
0 00-
PF>m 2ndHDM-»-
M-7 = 0.05 ppm
024681
Day
Figure 9-16. Diagrammatic representation of several exposure indices, illustrating how
they weight concentration and accumulate exposure, (a) SUM06 - The
upper graphic illustrates an episodic exposure profile; the shaded area under
some of the peaks illustrates the concentrations greater than or equal to 0.06
ppm that are accumulated in the index. The insert shows the concentration
weighting (0 or 1) function. The lower portion of the graphic illustrates how
concentration is accumulated over the exposure period, (b) SIGMOID - The
upper graphic illustrates an episodic exposure profile; the variable shaded
area under the peaks illustrates the concentration dependent weights that are
accumulated in the index. The insert shows the sigmoid concentration
weighting function. The mid-point of the sigmoid weighting scheme was
0.062 ppm. The lower portion of the graphic illustrates how concentration is
accumulated over the exposure period, (c) 2ndHDM and M-7 - The upper
graphic illustrates an episodic exposure profile. The lower portion of the
graphic illustrates that the 2ndHDM considers only a single exposure peak
while the mean applies a constant exposure value over the exposure period.
Source: Tinge etal. (1991).
1 Environmental Protection Agency, 1986). These studies previously reviewed by EPA (U.S.
2 Environmental Protection Agency, 1992, 1996a) and were in general agreement and consistently
3 favored the use of cumulative peak-weighted exposure indices. Lee et al. (1987) suggested that
4 exposure indices that included all the data (24 h) performed better than those that used only 7 h
5 of data; this is consistent with the conclusions of Heagle et al. (1987) that plants receiving
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1 exposures for an additional 5-h/day showed 10% greater yield loss than those exposed for
2 7-h/day. In a subsequent analysis using more of the NCLAN data, Lee et al. (1988) found the
3 "best" exposure index was a phenologically weighted, cumulative index, with sigmoid weighting
4 on concentration and a gamma weighting function as a surrogate for plant growth stage. This
5 index was the best statistical fit, but it depended upon more knowledge of species and site
6 conditions that made specification of weighting functions difficult for general use. This type of
7 multi-function index did not meet the other criteria for developing an air quality standard, that
8 being simplicity and understandability (Blanchard and Farley, 1999).
9 The next best fits were the several indices which only cumulated and weighted higher
10 concentrations (e.g., SUM06, SUM08, AOT40, W126). Amongst this group it was not possible
11 to distinguish a single best fit (Lee et al., 1988). Similarly Lefohn et al., (1992) reported that it
12 was not possible to differentiate among the SUMOO, SUM06, SUM08, and W126 exposure
13 indices because the indices were highly correlated with one another in the experiment. Others
14 have reported similar results when attempting to identify optimum exposure indices (Musselman
15 etal., 1988).
16 Other factors, including predisposition time (Hogsett et al., 1988; McCool et al., 1988) and
17 crop development stage (Heagle et al., 1991; Tingey et al., 2002), contribute to variation in the
18 biological response and suggest the need for weighting O3 concentrations to account for
19 predisposition time and phenology. However, the roles of predisposition and phenology in
20 influencing plant response vary considerably with species and environmental conditions, so that
21 specification of a weighting function for general use in characterizing plant exposure is not
22 possible at this time.
23 In similar retrospective analyses, using data from the European Open-Top Chamber
24 Program, Finnan et al. (1997) confirmed that cumulative exposure indices which emphasize
25 higher O3 concentrations are best related to plant response and that cumulative exposure indices
26 which use weighting approaches provide a better fit (U.S. Environmental Protection Agency,
27 1996a).
28 The main conclusions from 1996 regarding a biologically-relevant index based on ambient
29 exposure still hold true today. No information has come forth in the interim that alters those
30 conclusions, and in fact, some recent studies have further substantiated them. These key
31 conclusions can be restated as follows:
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I • ozone effects are cumulative;
2 • peak concentrations appear to be more important than lower concentrations in eliciting
a response;
3 • plant sensitivity to O3 varies with time of day and crop development stage;
4 • exposure indices that accumulate the O3 exposure and preferentially weight the peaks
yield better statistical fits to response than do the mean and peak indices;
I • to avoid possible overestimation of vegetation effects when using NCLAN-type yield
data in exposure-response models, the cumulative exposure index might be combined
with number of hours of high hourly average concentrations over the time period.
2 Following the 1996 review process (U.S. Environmental Protection Agency, 1996a,b), the
3 EPA proposed an alternative form of the secondary NAAQS using a cumulative, concentration-
4 weighted exposure index, the SUM06, to protect vegetation from damage (Federal Register,
5 1997). The EPA considered three specific concentration-weighted indices: the threshold
6 weighted SUM06, the AOT40, and the sigmoid weighted W126 exposure index (U.S.
7 Environmental Protection Agency, 1996b). Both indices performed equally well in predicting
8 the exposure-response relationships observed in the crop and tree seedlings studies conducted
9 during the prior 20 years (Heck and Cowling, 1997). In the absence of research results that
10 differentiate the predictive power of these two forms, the EPA selected the SUM06 exposure
11 index as the form for the proposed secondary standard, recognizing its simplicity,
12 understandability, and ease of use (U.S. Environmental Protection Agency, 1996b). A 3-month,
13 12-hour SUM06 exposure index of 26.4 ppm hr was proposed (U.S. Environmental Protection
14 Agency, 1996b). The value represented the concentration level that would protect 50% of the
15 crop species from a 10% yield loss. The composite response model included 49 studies
16 comprising 13 crop species and 5 locations across the U.S. where the crops are grown (U.S.
17 Environmental Protection Agency, 1996a).
18 European scientists took a similar approach as the United States in developing indices
19 describing growth and yield loss in crops and tree seedlings, using open-top chambers with
20 modified ambient exposures, but many fewer crop species were employed in the European
21 studies. The European countries were seeking a scientific basis for control strategies to reduce
22 air pollution. They adopted the critical levels and loads approach, as per the UN Economic
23 Commission for Europe (UNECE). A critical level was defined as "the concentration of
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1 pollutant in the atmosphere above which direct adverse effects on receptors, such as plants,
2 ecosystems, or materials may occur according to present knowledge" (United Nations Economic
3 Commission for Europe, 1988). Critical levels are set to prevent long-term injury and damage to
4 the most sensitive elements of any ecosystem. They are used to map and identify areas in
5 Europe in which the levels are exceeded and that information is then used to plan optimized and
6 effect-based abatement strategies. As used by the UNECE, they are not air quality standards in
7 the US sense, but they have been used as targets for planning reductions in sulfur and nitrogen
8 emissions to protect ecological resources. The nature of the significant harmful effects is not
9 specified in the original definition, which provides for different levels for different types of
10 harmful effect (e.g., visible injury or loss of crop yield). There are also different levels for crops,
11 forests, and semi-natural vegetation. The caveat, "according to present knowledge," is important
12 because critical levels are not rigid; they are revised periodically as new scientific information
13 becomes available. For example, the original critical level for O3 specified concentrations for
14 three averaging times, but further research and debate led to the current critical levels being
15 stated as the cumulative exposure (concentration x hours) over a cut-off concentration of 40 ppb
16 (AOT40) (Fuhrer et al., 1997). The "Level I" critical level was used in the 1990s to map areas of
17 exceedance, but analyses of many exposure studies led to the conclusion that the simple,
18 exposure-based level leads to over-estimation of the effects in some regions and
19 under-estimation in others (Fuhrer et al., 1997; Karenlampi and Skarby, 1996). The main
20 problem was that other environmental factors (vapor pressure deficit, water stress, temperature,
21 and light) and variation in canopy height altered O3 uptake and its effects.
22 A decision was made to work towards a flux-based approach for the critical level, with the
23 goal of modeling O3 flux-effect relationships for three vegetation types: crops, forests, and
24 semi-natural vegetation. Progress has been made in modeling flux (e.g., Grunhage and Jager,
25 2003; Ashmore et al., 2004a; Ashmore et al., 2004b) and the Mapping Manual is being revised
26 (Karlsson et al., 2003; Ashmore et al., 2004a; Ashmore et al., 2004b; Grennfelt, 2004). The
27 revisions may include a flux-based approach for 3 crops: wheat, potatoes, and cotton; but,
28 because of lack of flux data, a cumulative, concentration-based (AOTx) exposure index will
29 remain for most crops, and for forests and semi-natural herbaceous vegetation (Ashmore et al.,
30 2004; Jaeger and Grunhage, 2004).
31
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1 9.5.4 Identifying Exposure Components That Relate to Vegetation Effects
2 The efficacy of exposure indices to predict biological response requires that researchers
3 identify a relationship between exposure components and effects, as well as those environmental
4 and site factors that control pollutant uptake by the plant. These relationships were identified
5 and discussed in the 1996 review (U.S. Environmental Protection Agency, 1996a).
6 A significant, but in some instances, unquantifed role was identified for: (1) concentration;
7 (2) duration of exposure; (3) the diurnal and seasonal patterns of exposure, e.g., time of day of
8 peak event, season of higher exposures, seasons of high precipitation and humidity, the
9 frequency of occurrence of peak events and respite time (peak to valley ratios); (4) plant
10 phenology; (5) plant canopy structure; (6) meteorological and site factors, e.g., light, humidity;
11 and (7) plant defense mechanisms.
12
13 9.5.4.1 Role of Concentration
14 A significant role of higher concentrations was established earlier, based on several
15 experimental studies (U.S. Environmental Protections Agency, 1996a). Recently Nussbaum
16 et al., (1995) and Yun and Laurence (1999b) have added support for the important role that peak
17 concentrations, as well as the pattern of occurrence, plays in plant response to O3. Based on air
18 quality data from 10 U.S. cities, three treatments of 4-week exposure to the same SUM06 value
19 were constructed by Yun and Laurence (1999b). They used the regimes to explore effects of
20 treatments with variable peak occurrence versus uniform peak occurrence during the exposure
21 period. The authors reported that the peak exposures were important and that the same SUM06
22 resulted in very different foliar injury. Oksanen and Holopainen (2001) found that the peak
23 concentrations and the shape of the O3 exposure (i.e., duration of the event) were important in
24 foliar injury of white birch saplings, but growth reductions were found to be more related to
25 cumulative exposure. Nussbaum et al. (1995) also found peak concentrations and the pattern of
26 occurrence to be critical in the measured response. The authors recommended that to describe
27 the effect on total forage yield, peak concentrations > 0.11 ppm must be emphasized by using an
28 AOT with higher threshold concentrations.
29 A greater role for higher concentrations affecting plant growth may be inferred based on
30 recent air quality analyses for the Southern California area (Lee et al., 2003; Tingey et al., 2004).
31 In the late 1960s and 1970s, extremely high O3 concentrations had impacted the San Bernardino
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1 National Forest. However, over the past 15 plus years, significant reductions in the O3 exposure
2 have occurred in the San Bernardino National Forest (Lee et al., 2003; Lefohn and Shadwick,
3 2000; Davidson, 1993; Lloyd et al., 1989). An illustration of this improvement in air quality is
4 shown by the 37 year history of ozone air quality at a site in the San Bernardino Mountains
5 (Figure 9-17) (Lee et al., 2003). The O3 exposure increased from 1963 to 1979 concurrent with
6 increased population and vehicular miles, followed by a decline to present mirroring decreases in
7 precursor emissions. The pattern in exposure was evident in various exposure indices including
8 the cumulative concentration weighted (SUM06), as well as maximum peak event (1-h peak),
9 and the number of days having hourly averaged O3 concentrations > 95 ppb (i.e., the California
10 ozone standard). The number of days having hourly averaged O3 concentrations > 95 ppb
11 declined significantly from 163 days in 1978 to 103 days in 1997. The changes in ambient
12 ozone air quality for the site were reflected in the changes in the frequency and magnitude of the
13 peak hourly concentration and the duration of the exposure (Figure 9-17). Considering the role
14 of exposure patterns in determining response, the seasonal and diurnal patterns in hourly O3
15 concentration did not vary appreciably from year to year over the 37-year period (Lee et al.,
16 2003).
17 The inference for a role of higher concentrations comes from both results of ground
18 measures of tree conditions on established plots and from model simulations. Across a broad
19 area of the San Bernardino National Forest, the Forest Pest Management (FPM) method of injury
20 assessment indicated an improvement of crown condition from 1974 to 1988; and the area of
21 improvement in injury assessment is coincident with an improvement of O3 air quality (Miller
22 and Rechel, 1999). A more recent analysis of forest changes in the San Bernardino National
23 Forest using an expanded network of monitoring sites has verified significant changes in growth,
24 mortality rates, basal area, and species composition throughout the area since 1974 (Arbaugh
25 et al., 2003). A model simulation of ponderosa pine growth over the 40 year period in the
26 San Bernardino showed a significant impact of ozone exposure on tree growth and indicates
27 improved growth with improving air quality. The improvement in growth was assigned to
28 improved air quality, but no distinction was made regarding the relative role of mid-range and
29 higher hourly concentrations, only that improved growth tracked both decreasing SUM06,
30 maximum peak concentration and number of days of hourly O3 > 95 ppb (Tingey et al., 2004).
31
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1 9.5.4.2 Role of Duration
2 Recent studies have called into question the period of time over which concentrations are
3 accumulated and the form of the exposure index. Heagle and Stefanski (2000) reported that the
4 form of the exposure index was important only for 24-h indices for which SUMOO provided the
5 poorest fit. The authors reported that the SUMOO, SUM06, W95 (Lefohn and Runeckles, 1987),
6 W126, and AOT40 produced similarly good fits of the foliage biomass data for 6-, 5-, and 4-h
7 midday accumulating periods. The study "pooled" data from San Bernardino (CA) and
8 Riverside (CA) with data from Amherst (MA), Corvallis (OR), Kennedy Space Center (FL),
9 Raleigh (NC), and Blacksburg (VA). Ozone exposures were much higher at the two California
10 sites (indicated by high W126, SUM06, W95, and AOT40 values) in comparison to the other
11 locations. Because of the pooling of the data, the large number of high hourly average O3
12 concentrations that occurred at the California sites may have resulted in the exposure indices
13 being highly correlated with one another and made it difficult to identify one optimal index.
14 In another study in California, Arbaugh et al. (1998) reported that the SUMOO exposure
15 index performed better for describing visible injury than the SUM06, W126, number of hours
16 greater than or equal to 0.08 ppm, and the number of days between measurement periods. These
17 exposure indices were originally developed and tested using only growth/yield data, not foliar
18 injury (U.S. Environmental Protection Agency, 1996a). This distinction is critical in comparing
19 the efficacy of one index to another.
20 The SUMOO exposure index is a surrogate for a long-term average concentration and
21 earlier growth response studies indicated that a long-term average was not adequate in predicting
22 effects primarily because it gives equal weight to all concentrations (U.S. Environmental
23 Protection Agency, 1996). However, for many locations in California, a large number of higher
24 hourly average concentrations occur and the SUMOO could be highly correlated with the
25 frequency of elevated hourly average concentrations and thus could be a good predictor of
26 vegetation effects. In Section 9.3, it was noted that research results are mixed on what causes an
27 effect in plants: an event or the cumulation of events.
28
29 9.5.4.3 Patterns of Exposure
30 A significant factor in developing exposure indices is the temporal patterns of ozone
31 occurrence over a day, a month, a year, and seasonally overlaying the daily and seasonal
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1 temporal patterns of those influential climatic and site factors. The coincidence of peak ozone
2 and maximal stomatal conductance and detoxification processes are key to effecting plant
3 growth response (Musselman and Minnick, 2000).
4
5 Daily Patterns
6 Recent experimental studies with tree species have demonstrated the de-coupling of
7 ambient O3 exposure, peak occurrence, and gas exchange, particularly in areas of drought
8 (Panek, 2004). The coincidence of maximal conductance and occurrence of higher
9 concentrations are relevant to the time of day that exposure is cumulated in those indices that
10 cumulate and weight concentration.
11 The hours during the day over which ambient exposure is cumulated or effective dose is
12 estimated is a surrogate for the coincidence of these patterns of conductance and O3 occurrence.
13 A 12-hr daylight period for cumulating exposure was proposed following the 1996 review (U.S.
14 Environmental Protection Agency, 1996). An extensive review of the literature revealed that a
15 large number of species had varying degrees of nocturnal stomatal conductance (Musselman
16 and Minnick, 2000). Grulke et al. (2003) showed that the stomatal conductance at night for
17 ponderosa pine in the San Bernardino (CA) National Forest ranged from one tenth to one fourth
18 that of maximum daytime gas exchange. In June, at the high-elevation site, 11% of the total
19 daily O3 uptake of pole-sized trees occurred at night. In late summer, however, O3 uptake at
20 night was negligible. Birch seedlings exposed to O3 at night show greater reductions in growth
21 than those exposed to O3 in daylight (Matyssek et al., 1995). Brassica rapa plants exposed to O3
22 during the day or night, show little significant difference in the amounts of injury or reduced
23 growth response to treatment; the conductance was 70 to 80% lower at night (Winner et al.,
24 1989). Tissue biomass of ponderosa pine seedlings was significantly reduced when seedlings
25 were exposed to either daytime or nighttime episodic profiles (Lee and Hogsett 1999). However,
26 the biomass reductions were much greater with daytime peak concentrations than with nighttime
27 peak concentrations.
28 Although conductance is lower at night than during the day for most plants, nocturnal
29 conductance can result in some measurable O3 flux into the plants and should be considered.
30 Nocturnal O3 flux also depends on the level of turbulence that intermittently occurs at night.
31 In addition, plants can be more susceptible to O3 exposure at night than during the daytime,
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1 because of lower plant defenses at night (Musselman and Minnick, 2000). Massman (2004)
2 suggested that nocturnal stomatal O3 uptake accounted for about 15% of the cumulative daily
3 effective O3 dose that was related to predicted injury. Based on a review of the literature relating
4 to nocturnal stomatal conductance, Musselman and Minnick (2000) recommended that any O3
5 exposure index used to relate air quality to plant response should use the 24-h cumulative
6 exposure period for both exposure-response and effective flux models.
7 In general, stomatal conductance needs to be taken into account in developing indices
8 (Panek et al., 2002). Stomatal conductances are linked to both diurnal and seasonal
9 meteorological activity and soil/site conditions (e.g., soil moisture). Daily patterns of
10 conductance are often highest in mid-morning, whereas higher ambient O3 concentrations
11 generally occur in mid to late afternoon, when stomata are often partially closed and
12 conductances are lower. Total flux depends on atmospheric resistance and boundary layer
13 resistances, both of which exhibit variability throughout the day. Several recent studies have
14 suggested that ponderosa pine trees in the southern and northern Sierra Nevada Mountains may
15 not be as sensitive to high O3 concentrations as to lower concentrations, due to reduced O3
16 uptake during the period when the highest concentrations occur (Panek et al., 2002; Panek and
17 Goldstein 2001; Bauer et al., 2000; Arbaugh et al.,1998). Panek et al. (2002) compared direct
18 measurements of ozone flux into a canopy of ponderosa pine and demonstrated a lack of
19 correlation of daily patterns of conductance and ozone occurrence, especially in the late season
20 drought period; and they concluded that a consideration of climate or season was essential,
21 especially considering the role of soil moisture and conductance/uptake. In contrast, Grulke
22 et al. (2002) reported high conductance when O3 concentrations were high in the same species,
23 but different growing site conditions. The uncoupling of conductance and higher ambient O3
24 concentration would hold true for more mesic environments as well as xeric landscapes. The
25 longer term biological responses reported by Miller and Rechel (1999) for ponderosa pine in the
26 same region and the general reduction in recent years in ambient O3 concentrations, suggest that
27 conductance alone may not be a sufficient indicator of potential vegetation injury or damage.
28 The generalized models of stomatal conductance may provide a means to link patterns of
29 O3 occurrence with climatic and site factors that affect O3 uptake to some degree, provided
30 conductance is modeled by regions of similar seasonal moisture and by similar canopy structure
31 (e.g., Emberson et al., 2000a,b; Grunhage et al., 2000; Massman, 2004)
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1 Seasonal Patterns
2 Several of the recent studies measuring O3 flux to pine canopies also reported on the
3 importance of seasonal patterns in relating exposure to response (Panek, 2004). These seasonal
4 patterns can be early vs late season occurrence of higher O3 concentrations, reflecting climate
5 and precursor availability. The patterns also reflect seasonal drought and the role of soil
6 moisture plays in stomatal conductance and O3 uptake. Recently, studies have looked directly at
7 this linkage. Panek et al. (2002) compared direct measurements of ozone flux into a canopy of
8 ponderosa pine with a number of exposure indices, demonstrated a lack of correlation especially
9 in the late season drought period, and concluded that a consideration of climate was essential,
10 especially soil moisture. They suggested that a better metric for a seasonally drought-stressed
11 forests would be one that incorporates forest physiological activity, either through mechanistic
12 modeling, by weighting ambient O3 concentrations by stomatal conductance, or by weighting O3
13 concentrations by site moisture conditions. Panek (2004) demonstrated a decoupling of O3
14 exposure and uptake seasonally as well, via seasonal drought influence. Maximum O3 uptake
15 occurred at the beginning of the season and in the winter, whereas the pines were nearly dormant
16 during August-September.
17 Using TREGRO, a process-based model, Tingey et al. (2004) simulated long-term growth
18 of ponderosa pine over a 37-year period. The simulation showed high correlation between O3
19 exposure and O3-induced reductions in tree growth (R2 = 0.56). The scatter about the line
20 however indicates that other factors besides O3 are required to describe the association between
21 exposure and response. Incorporating annual temperature and precipitation increased the R2 to
22 0.67. In keeping with the observations of Panek (2004) on the decoupling of peak O3 occurrence
23 and maximal conductance, the remaining unexplained variation is attributed to differences in
24 timing of peak O3 uptake and peak O3 exposure over the years.
25
26 9.5.4.4 Frequency of Occurrence of Peak Concentrations
27 Several earlier studies demonstrated the greater effect of episodic occurrence of O3 peaks
28 compared to daily peak events (U.S. Environmental Protection Agency, 1996a); and, since the
29 last review, a few studies have corroborated the importance of this pattern in growth response.
30 Kollner and Krause (2003) reported that, under equal exposure conditions, the most pronounced
31 effects on the yield of sugar beet and soybeans occurred with those regimes that emphasized the
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1 episodic occurrence of peak events. Nussbaum et al. (1995) compared the effects of different
2 patterns of peak occurrences with similar AOT40 values and reported a stronger effect on total
3 forage yield from the episodic treatment.
4
5 9.5.4.5 Canopy Structure
6 Another factor important in either O3 exposure or uptake, is how canopy structure affects
7 O3 concentration in and under forest canopies. There have been several comprehensive studies
8 of O3 concentrations under tree canopies (Enders, 1992; Fontan et al., 1992; Fredericksen et al.,
9 1995; Joss and Graber, 1996; Kolb et al., 1997; Lorenzini and Nali, 1995; Neufeld et al., 1992;
10 Samuelson and Kelly, 1997). In general they indicate a reduction in O3 of-20 to 40% below the
11 canopy but above the shrub/herb layers. An essential component in the critical level AOT40 is
12 the height of where the O3 concentration is measured. The critical levels are related to the O3
13 concentration measured at the top of the canopy, i.e., upper surface boundary of the (quasi-)
14 laminar layer (Grunhage et al., 2003). This location is presumably more related to stomatal
15 uptake. Weighting the O3 concentration at this location takes into account stomatal opening and,
16 if weighted with the Jarvis-Steward factors for radiation, temperature, and soil moisture, the
17 "lexicologically" effective AOT40 is obtained (Grunhage et al., 2003).
18 In a study that considers those factors important in O3 uptake that are also spatially
19 distributed as a result of canopy structure, Davison et al. (2003) reported that the variation in
20 visible injury in coneflower populations is unlikely to be due to differences in O3 flux and more
21 likely due to variation in PAR. At a height of 50 cm above ground, PAR was reduced by almost
22 90%, whereas the O3 varied from about 15 to 90% of ambient. Ozone injury was not solely
23 related to O3 flux. Although there have been studies of the effects of different light levels on O3
24 response, there have been few at the very low levels that occur in canopies of tall herbaceous
25 stands or in the ground layer of forests. Davison et al. (2003) report that conductance was not
26 related to diurnal changes in light. The O3 levels were still about 90% of the O3 concentration
27 above the canopy when light was less than 5%. Light intensity dropped to 1.5% of open at 130
28 cm from the edge of the canopy, while O3 dropped to only 42%. The study, although reporting
29 on the adequacy of visible foliar injury as an indicator of O3 effects, does suggest that
30 consideration of other factors such as light are important in predicting response. How this may
31 be included in developing exposure-response indices was not considered.
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1 9.5.4.6 Site and Climate Factors
2 Soil moisture is a critical factor in controlling O3 uptake through it's effect on plant water
3 status and stomatal conductance.. In an attempt to relate uptake, soil moisture, and ambient air
4 quality to identify areas of potential risk, available O3 monitoring data for 1983 to 1990 were
5 used, along with literature-based seedling exposure-response data from regions within the
6 southern Appalachian Mountains that might have experienced O3 exposures sufficient to inhibit
7 growth (Lefohn et al., 1997). In a small number of areas within the region, O3 exposures and
8 soil moisture availability were sufficient to possibly result in growth losses in some sensitive
9 species (e.g., black cherry). The conclusions were limited because of the interpolation of the O3
10 exposures in many of the areas and the hydrologic index used might not reflect actual water
11 stress.
12
13 9.5.4.7 Plant Defense Mechanism - Detoxification
14 The non-stomatal component of plant defenses are the most difficult to quantify, but some
15 studies are available (Barnes et al., 2002; Plochl et al., 2000; Chen et al., 1998; Massman and
16 Grantz, 1995). Massman et al. (2000) developed a conceptual model of a dose-based index to
17 determine plant injury response to O3 that relates to the traditional exposure-based parameters.
18 The index uses time-varying-weighted fluxes to account for the fact that flux is not necessarily
19 correlated with plant injury or damage. Their model applies to plant foliar injury, and they
20 suggest that application of flux-based models for determining plant damage (yield or biomass)
21 will require a better understanding and quantification of the injury and damage relationship.
22
23 9.5.5 Ozone Uptake or Effective Dose as an Index
24 Developing an index that relates growth response to ambient exposure has been
25 approached in the past through various weighting functions on those ambient exposure factors,
26 including concentration, duration, and time of day for cumulating exposure. These indices have
27 not incorporated factors of climate patterns and species- and site factors that control O3 uptake
28 via canopy and stomatal conductance. The other approach is basing the index on the O3
29 concentration going into the leaf, or flux. This approach includes those factors controlling
30 uptake through canopy and stomatal conductance and, by necessity, relies on models to predict
31 flux and ultimately the "effective" flux: that concentration that reaches the target tissue to cause
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1 an effect (Grunhage et al., 2004; Massman 2004; Massman et al., 2000) .Effective flux has been
2 defined as the balance between the O3 flux and the detoxification process (Dammgen et al.,
3 1993; Grunhage and Haenel, 1997; Musselman and Massman, 1999). The time-integrated
4 effective flux is termed "effective dose". The uptake mechanisms and the resistances in this
5 process, including stomatal conductance and biochemical defense mechanisms, are discussed in
6 the previous Section 9.3.
7
8 9.5.5.1 Models of Stomatal Conductance
9 Only a limited number of studies have measured O3 concentration or its reaction products
10 within the leaf. Otherwise the index of uptake is a modeled result that considers site, climatic,
11 meteorological, and species-specific (e.g., detoxification reactions) factors. Models of O3
12 conductance into plant tissue are available (Grunhage and Jager, 2003; Emberson et al., 2000a,b;
13 Grunhage and Haenel, 1997; Grunhage et al., 1997; Massman, 1993; Wesely, 1989). The
14 European Monitoring and Evaluation Program (EMEP) has an O3 deposition model developed
15 for application across Europe in conjunction with the EMEP photochemical model as a tool for
16 the critical levels program. The model has been developed to estimate vegetation type-specific
17 O3 deposition and stomatal flux, calculated according to a standard 3-resistance formulation
18 incorporating atmospheric, boundary layer, and stomatal resistances (Emberson 2000b). The
19 model uses a multiplicative algorithm of the stomatal conductance of O3 (Jarvis, 1976) and has
20 been parameterized for 10 European tree species, 7 agricultural species, and 1 type of semi -
21 natural vegetation. The model calculates conductance as a function of leaf phenology,
22 temperature, photosynthetic flux density (PFD), vapor pressure deficit (VPD), and soil moisture
23 (SMD). The environmental variables are site-specific (or regionally-specific). The most
24 important factors limiting O3 with this model were vapor pressure deficit (VPD), soil moisture
25 deficit, and phenology (Emberson et al., 2000b). These factors demonstrate the critical linkage
26 of high VPD and stomatal closure, which typically co-occurs with high O3 concentrations.
27 A number of recent model-based studies relating flux and plant growth response have
28 investigated several crop and forest tree species (Karlsson et al., 2004; Karlsson et al., 2004;
29 Pleijel et al., 2004; Matyssek et al., 2004; Elvira et al., 2004; Soja et al., 2004; Weiser and
30 Emberson, 2004; Touvinen et al., 2004; Altimir, et al., 2004; Gerosa et al., 2004; Mikkelsen,
31 et al., 2004, Bassin, et al., 2004; Emberson et al., 2000). The studies have used earlier exposure
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1 experiments as well as explicitly designed field studies but, to date, no abundance of data that
2 would provide the basis for a flux-based index.
3 The complexity of using flux as an index of O3 exposure for response is shown in field
4 studies that measured O3 flux into Norway spruce and cembran pine (Wieser et al., 2000). They
5 demonstrated that stomatal conductance was the main limiting factor for O3 uptake and showed
6 the dependence of that measure on crown position, needle age, and altitude. Consideration of the
7 role of climate illustrates the importance of a flux measure. Pleijel et al., (2000) reported the
8 improved relationship of yield in spring and winter wheat grown in OTCs in many areas across
9 Europe when it was related to the cumulative stomatal O3 uptake during the grain-filling period.
10 Compared to the AOT40, the cumulative uptake index estimated larger yield losses in the
11 relatively humid parts of western and northern Europe, while smaller yield loss was estimated
12 for the dry summer climates in south and central Europe.
13 Danielsson et al. (2003) compared the ability of two different stomatal models to relate
14 grain yield in field grown spring wheat to cumulated O3 uptake and an exposure index of
15 AOT40, and found that the cumulated O3 uptake determined with either model performed better
16 in relating exposure to yield than did the cumulative exposure index of AOT40.
17 Cumulative O3 uptake was modeled for three deciduous and 2 confierous species growing a
18 different sites and elevations and compared with exposure measure of AOT40 at these sites
19 (Matyssek et al., 2004). A general linearity was demonstrated between the two measures of O3
20 exposure; and, at any given AOT40, there was a 25 ± 11% variation in CU. Although no
21 correlation of growth alterations was observed with either the exposure or the uptake measure,
22 the modeled cumulative uptake was able to describe the variation in tree size and site location
23 which makes for a better measure in risk assessment of O3 (Matyssek et al., 2004). Karlsson
24 et al. (2004) compared the biomass-response relationship in young trees at seven experimental
25 sites across Europe using modeled cumulative O3 uptake and AOT40. A weaker dose-response
26 relationships were reported for the cumulative uptake metric than the AOT40 (Karlsson et al.,
27 2004),
28 Concern about the complexity of the stomatal models and the data needed to model O3
29 uptake, has led some researchers to offer modified accumulated exposure indices that consider
30 those meteorological factors controlling uptake (Karlsson et al., 2004; Gerosa et al., 2004). In a
31 study of subterranean clover in Austria, Belgium, and southern Sweden, Karlsson et al., (2004)
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1 reported on the performance of a modified accumulated exposure over the threshold (mAOT)
2 which was based on solar radiation and vapor pressure deficit. This index improved the
3 relationship for observed visible injury. But if modeled uptake of O3 was derived from a simple
4 stomatal conductance model considering solar radiation, VPD, and air temperature, then this
5 index gave an even greater improvement in the relationship to visible injury than did the ambient
6 exposure index of AOT40 (Karlsson et al., 2004B). The added value of the mAOT was
7 worthwhile and it had a lower degree of complexity and data requirements modeling O3 uptake
8 with stomatal models. Based on a study of O3 fluxes over a barley field in Italy, a similar
9 modified exposure index was reported and referred to as "effective exposure" (Gerosa et al.,
10 2004). Their approach was similar in its consideration of physiological aspects in conjunction
11 with monitored O3 concentrations. It addressed the shortcoming of the data needs for modeled
12 O3 uptake.
13 Models that partition O3 uptake into stomatal and non-stomatal components are also now
14 available and predict a significant non-stomatal component in calculating O3 flux (Altimir et al.,
15 2004; Mikkelsen et al., 2004; Nikolov and Zeller, 2003; Zeller and Nikolov, 2000; Bassin et al.,
16 2004; Nussbaum et al., 2003). Altimir et al. (2004) compared the relative contribution of
17 stomatal and non-stomatal sinks at the shoot level for Scots pine. Using the EMEP model with a
18 revised parameterization for Scots pine, they demonstrated that a major removal of O3 was due
19 to the non-stomatal component; and when a non-stomatal term was introduced dependent on
20 ambient relative humidity, the non-stomatal contribution to the total conductance was about
21 50%. Zeller and Nikolov (2000) demonstrated a large non-stomatal O3 uptake (41% of the total
22 annual flux) in subalpine fir at a site in southern Wyoming using the biophysical model
23 FORFLUX. In a 5 year study of measured O3 flux to a Norway spruce canopy, Mikkelsen et al.
24 showed monthly patterns of non-stomatal and stomatal deposition as part of total deposition to
25 the canopy. Their study demonstrated that daily means of O3 concentration and fluxes averaged
26 over 5 years correlate well, but the correlation is based on two different uncoupled processes
27 outside and inside the stomates. The destruction of O3 in the canopy outside the stomates is
28 influenced by temperature, light and humidity, e.g., surface reactions, NO- and VOC-emissions;
29 and these same factors influence stomatal opening, e.g., mid-day and night closure.
30 Consequently, the diurnal O3 concentration and O3 flux do not correlate at all during the growing
31 season. The study estimated yearly stomatal uptake to be a minimum of 21% of total deposition
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1 (i.e., non-stomatal as high as 80% of total). The stomatal uptake is highest in May-August (30-
2 33%) and lowest in November-February (4-9%).
3
4 9.5.5.2 Nonlinear Response and Developing Flux Indices
5 If only O3 flux were used as a metric to predict vegetation injury or damage, the prediction
6 might be overestimated because of reported non-linear relationships between O3 and plant
7 response (Amiro et al., 1984; Amiro and Gillespie, 1985; Bennett, 1979; U.S. Environmental
8 Protection Agency, 1978; 1986; 1996). The non-linearity in the response surface suggested the
9 existence of a biochemical threshold. More recently, nonlinear relationships between O3 flux
10 and yield were shown for potato (Pleijel et al.,2002) and spring wheat (Danielsson et al.,2003).
11 The relationship between O3 flux and potato yield suggested using an instantaneous flux
12 threshold to overcome the nonlinear relationship (Pleijel et al., 2002). However, the authors did
13 not report a substantial improvement in the mathematical fitting of the model when applying the
14 threshold Most of the flux was accumulated below 0.06 ppm. Danielsson et al. (2003) was able
15 to show an improved relationship between O3 uptake and yield of spring wheat using a threshold
16 of 5 nmoles m"2 sec"1 (0.24 mg m"2 sec"1). These results suggest not all O3 entering the stomata
17 contribute to a reduction in yield, which depends to some degree on the amount of internal
18 detoxification occurring with each particular species (see Section 9.3). The fact that the defense
19 and repair mechanisms vary diurnally as well as seasonally may make it extremely difficult to
20 apply a mathematically determined threshold to instantaneous flux measurements to calculate
21 cumulative flux. The threshold models do not allow for the temporal (i.e., daily and seasonal)
22 variability of defense mechanisms.
23 Musselman and Massman (1999) suggest that those species having high amounts of
24 detoxification potential may show less of a relationship of O3 stomatal uptake to plant response.
25 There has been no direct experimental demonstration of this relationship, however. The cellular
26 detoxification reactions and repair processes which both detoxify oxidants as well as play central
27 role in the carbon economy of the plant are another level of resistance to O3 reaching the target
28 tissue (see Section 9.3). As indicated earlier, effects occur on vegetation when the amount of
29 pollutant absorbed exceeds the ability of the plant to detoxify O3 or to repair the initial impact.
30 The magnitude of the response is determined by the amount of the pollutant reaching the target
31 site and the ability of the plant to reestablish homeostatic equilibrium. Thus, one would expect
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1 to observe a decoupling of O3 uptake with vegetation effects, which would manifest itself as a
2 nonlinear relationship between O3 flux and injury or damage.
3 Additional factors for inclusion in flux-based models to predict vegetation effects are the
4 defense and repair mechanisms. Specifically, the relationship between conductance, O3
5 concentration, and defense/repair mechanisms needs to be included. Recently, Massman (2004)
6 described results that illustrate that the combination of conductance, O3 concentration, and
7 diurnal variation of defense mechanisms show the daily maximum potential for plant injury
8 (based on effective dose) coincides with the daily peak in O3 mixing ratio. Massman et al.
9 (2000) stress that the product of the overlapping mathematical relationships of conductance,
10 concentration, and defense mechanisms results in a much different picture of potential impact to
11 vegetation than just the use of conductance and concentration in predicting vegetation effects.
12
13 9.5.5.3 Simulation Models
14 Another approach for determining O3 uptake and relating growth response to ambient O3
15 exposure may be the use of physiologically-based simulation models. Several of these have
16 been used in various contexts, comparing O3 response in a number of tree species with varying
17 climate and site factors (e.g., soil moisture) (e.g., Laurence et al., 2001; Ollinger et al., 1997;
18 Ollinger et al., 1998; Weinstein et al., 2001; Weinstein et al., 2004; Tingey et al., 2000; Tingey
19 et al., 2004). One of the important considerations in applying simulation modeling is to
20 carefully assess the uncertainties associated with the modeling predictions. Further efforts need
21 to be made to exercise the models so that they predict past growth losses associated with changes
22 in O3 exposures that can be verified with on-the-ground surveys.
23
24 9.5.6 Summary
25 A few studies published since 1996 have substantiated earlier conclusions on the role of O3
26 exposure components (including concentration, duration and exposure patterns) in describing
27 growth response to O3 exposures. Recent studies of different exposure patterns have confirmed
28 earlier studies on the importance of higher concentrations and duration of exposure when
29 describing response. An inferred role of peak concentrations is possible from consideration of
30 improved air quality in regions like the San Bernardino Mountains in southern California.
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1 Studies provide the basis for focusing on the higher O3 concentrations, while including the lower
2 levels, when estimating the effects of emission reductions on vegetation.
3 A few studies have demonstrated the potential disconnection of peak events and maximal
4 stomatal conductance. In addition, a few studies have demonstrated the uptake of O3 during
5 nighttime hours and suggested the need to cumulate O3 exposure 24 h per day and not just during
6 daylight hours.
7 Several studies since 1996 have demonstrated another critical concern in developing an
8 index for exposure. The concern is that peak O3 events and maximum stomatal conductance
9 may be temporally separate. This disconnection introduces uncertainty in assessing O3 impact
10 when using the current ambient exposure based cumulative, concentration weighted indices.
11 If stomatal conductance is relatively low, as in the late afternoon in arid climates, and that is the
12 same time as the peak O3 concentrations, then use of an exposure index that does not consider
13 this disconnect, will overestimate the effect of the exposure. This concern is especially apparent
14 when assessing the impact of O3 across all the varied climatic regions of the United States or
15 Europe. Some studies use stomatal models to predict uptake or using process level models (e.g.,
16 TREGRO) to integrate those climate and site factors that drive this temporal pattern of stomatal
17 conductance and exposure, and thus reduce some of the uncertainty in regional to national
18 assessments of effects. These approaches however are still species dependent.
19 The results of these studies and reviews have indicated the need to continue to develop
20 indices that are more physiologically and meteorologically connected to the actual dose of O3 the
21 plant receives. The cumulative concentration-weighted exposure indices are acknowledged
22 surrogates for effective dose that are simple conceptually and easy to measure. They do not fully
23 characterize the potential for plant uptake and resulting effects associated with O3 because the
24 indices, being measures of ambient concentration, do not include the physical, biological, and
25 meteorological processes controlling the transfer of O3 from the atmosphere through the leaf and
26 into the leaf interior. Use of such indices is especially limited in spatial risk characterizations
27 because of the lack of linkage of meteorology, species and site-specific factors influencing O3
28 uptake. The flux-based approach should provide an opportunity to improve upon the
29 concentration-based (i.e., exposure indices) approach. A cautionary argument was advanced in a
30 few publications centered around the non-linear relationship between ozone uptake and plant
31 injury (not growth alterations) response. The concern was that not all O3 stomatal uptake results
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1 in a reduction in yield, which depends to some degree on the amount of internal detoxification
2 occurring with each particular species. Those species having high amounts of detoxification
3 potential may show less of a relationship between O3 stomatal uptake and plant response.
4 The European approach and acceptance of flux-based critical values is a recognition of this
5 problem; and a concerted research effort is needed to develop the necessary experimental data
6 and modeling tools that will provide the scientific basis for such critical levels for O3 (Grunhage
7 et al., 2004; Fuhrer et al., 1997; Grunhage and Jager, 1994).
8 At this time, based on the current state of knowledge, exposure indices that differentially
9 weight the higher hourly average O3 concentrations, but include the mid-level values, represent
10 the best approach in the United States for relating vegetation effects to O3 exposure.. A large
11 database exists that has been used for establishing exposure-response relationships. Such a
12 database does not yet exist for relating O3 flux to growth response. Those pattern disconnects
13 between period of uptake and peak occurrence, as well as the potential for nocturnal uptake,
14 should be considered with some weighting functions in the currently used exposure indices.
15 Of particular consideration would be their inclusion in regional to national estimations of O3
16 impacts on vegetation. Another useful approach to regional assessment for given specie(s) is
17 simulating growth effects with process-based models that account for seasonal climate and site
18 factors that control conductance
19 It is anticipated that, as the overlapping mathematical relationships of conductance,
20 concentration, and defense mechanisms are better defined, O3 flux-based models may be able to
21 predict vegetation injury and/or damage at least for some categories of canopy-types with more
22 accuracy than the exposure-response models.
23
24
25 9.6 OZONE EXPOSURE-PLANT RESPONSE RELATIONSHIPS
26 9.6.1 Introduction
27 Ambient O3 concentrations have long been known to cause visible symptoms, decreases in
28 photosynthetic rates, decreases in plant growth, and decreases in the yield of marketable organs
29 (U.S. Environmental Protection Agency, 1978, 1986, 1996). Yet, despite considerable research
30 in the U.S. and other countries during the past three decades, quantifying the effects of ambient
31 O3 exposure on vegetation remains a challenge. Numerous studies have related O3 exposure to
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1 plant responses, with most effort focused on the yield of crops and the growth of tree seedlings.
2 Most experiments exposed individual plants grown in pots or soil under controlled conditions to
3 known concentrations of O3 for a segment of daylight hours for some portion of the plant's life
4 span (Section 9.2). The response of a plant species or variety to O3 exposure depends upon
5 many factors discussed in previous sections, including genetic characteristics (Section 9.4.2),
6 biochemical and physiological status (Section 9.4), and previous and current exposure to other
7 stressors (Sections 9.4, 9.5). Section 9.4 describes how O3 moves from the atmosphere into the
8 leaf and the subsequent biochemical and physiological responses of plants. The current section
9 focuses on the quantitative responses of plants to seasonal or multiyear exposures to known
10 amounts of O3. Quantitative responses include foliar symptoms and decreased growth of whole
11 plants or decreased harvestable portions of them. Because of the available information, most of
12 this section focuses on the response of individual plants, especially crop plants and tree
13 seedlings, with limited discussion of mixtures of herbaceous species. The responses of natural
14 ecosystems are discussed in Section 9.7.
15 This section will pay particular attention to studies conducted since the publication of the
16 1996 AQCD (U.S. Environmental Protection Agency, 1996). However, because much O3
17 research was conducted prior to the 1996 AQCD, the present discussion of vegetation response
18 to O3 exposure is largely based on the conclusions of the 1978, 1986, and 1996 criteria
19 documents (U.S. Environmental Protection Agency, 1978, 1986, 1996). To provide a context for
20 the discussion of recent research, the the key findings and conclusions of those three documents
21 are first summarized below.
22
23 9.6.2 Summary of Key Findings/Conclusions from Previous Criteria
24 Documents
25 Experimental data reviewed in the 1978 and 1986 criteria documents dealt primarily with
26 the effects of O3 on agricultural crop species (U.S. Environmental Protection Agency, 1978,
27 1986). The chapter on vegetation effects in the 1978 document (U.S. Environmental Protection
28 Agency, 1978) emphasized foliar symptoms and growth effects, but not those effects that
29 affected yield, an emphasis dictated by the kind of data available at the time. The 1986
30 document reviewed a substantial new body of evidence based on open-top chamber (OTC)
31 experiments (see Section 9.2) showing that ambient O3 exposures reduced the growth and yield
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1 of herbaceous plants, again with a focus on major crop species. In the 1986 and 1996
2 documents, data were presented from regression studies conducted to develop exposure-response
3 functions for estimating yield loss of major crop species in different regions of the United States.
4 The 1996 document included results from additional herbaceous crop species as well as shrub
5 and tree species. For a number of tree species, biomass growth of seedlings was related to
6 growing season O3 exposures to produce response functions for estimating O3 exposures that
7 reduce growth by 10 or 30%. Also, in the 1986 and 1996 documents, data from studies using
8 ethylene diurea (EDU) as a protectant were reviewed. The 1978, 1986, and 1996 criteria
9 documents also reviewed data on the response to O3 exposures of forest ecosystems in the
10 San Bernardino Mountains of southern California (U.S. Environmental Protection Agency, 1978,
11 1986, 1996). Because this region is exposed to high concentrations of O3 and has shown
12 evidence of ecosystem-level changes, it remains an important study area (see Section 9.7).
13 Ozone can cause a range of effects, beginning with individual cells, leaves, and plants, and
14 proceeding to plant populations and communities. These effects may be classified as either
15 injury or damage. Injury encompasses all plant reactions, such as reversible changes in plant
16 metabolism (e.g., altered photo synthetic rate), altered plant quality, or reduced growth that does
17 not impair yield or the intended use or value of the plant (Guderian, 1977). In contrast, damage
18 includes all effects that reduce or impair the intended use or value of the plant. Damage includes
19 reductions in aesthetic values as well as losses in terms of weight, number, or size of the plant
20 part that is harvested (yield loss). Yield loss also may include changes in crop quality, i.e.,
21 physical appearance, chemical composition, or the ability to withstand storage. Losses in
22 aesthetic values are difficult to quantify. Although foliar symptoms cannot always be classified
23 as damage, their occurrence indicates that phytotoxic concentrations of O3 are present, and,
24 therefore, studies should be conducted to assess the risk to vegetation.
25 Visible symptoms due to O3 exposures reduce the market value of certain crops and
26 ornamentals for which leaves are the product, e.g., spinach, petunia, geranium, and poinsettia.
27 The concept of limiting values used to summarize foliar symptoms in the 1978 document (U.S.
28 Environmental Protection Agency, 1978) was also considered valid in the 1986 document (U.S.
29 Environmental Protection Agency, 1986). Jacobson (1977) developed limiting values by
30 assessing the available scientific literature and identifying the lowest exposure
31 concentration/duration reported to cause foliar symptoms in a variety of plant species.
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1 A graphical analysis presented in those documents indicated that the limit for reduced plant
2 performance was an exposure to 50 ppb for several hours per day for more than 16 days.
3 Decreasing the exposure period to 10 days increased the concentration required to cause
4 symptoms to 100 ppb; and a short, 6-day exposure further increased the concentration to cause
5 symptoms to 300 ppb. These limiting values established in 1978 were still deemed appropriate
6 in the 1986 and 1996 criteria documents. Such foliar symptoms are caused by O3 concentrations
7 that occur in the United States as shown in Table 9-13 (as adapted from U.S. Environmental
8 Protection Agency, 1996).
9 The 1986 document emphasized that, although foliar symptoms on vegetation are often an
10 early and obvious manifestation of O3 exposure, O3 effects are not limited to foliar symptoms.
11 Other effects include reduced growth of many organs (including roots), changes in crop quality,
12 and alterations in plant susceptibility to biotic stressors and sensitivity to abiotic stressors. The
13 1986 document also emphasized that O3 exerts phytotoxic effects only if a sufficient amount of
14 O3 reaches sensitive sites within the leaf (Section 9.3). Ozone injury will not occur if the rate of
15 O3 uptake is low enough that the plant can detoxify or metabolize O3 or its metabolites or if the
16 plant is able to repair or compensate for the effects (Tingey and Taylor, Jr., 1982; U.S.
17 Environmental Protection Agency, 1986). Cellular disturbances that are not repaired or
18 compensated for are ultimately expressed as foliar symptoms, reduced root growth, or reduced
19 yield of fruits or seeds.
20 Beginning in the 1986 document and continuing in the 1996 document, OTC studies were
21 reviewed that better quantified the relationship between O3 exposure and effects on crop species,
22 with a focus on yield loss. These studies can be grouped into two types, depending on the
23 experimental design and statistical methods used to analyze the data: (1) studies that developed
24 predictive equations relating O3 exposure to plant response, and (2) studies that compared the
25 effects of discrete treatment level(s) to a control. The advantage of the regression approach is
26 that exposure-response models can be used to interpolate results between treatment levels.
27 Discrete treatment experiments were designed to test whether specific O3 treatments were
28 different from the control rather than to develop exposure-response equations, and the data were
29 analyzed using analyses of variance. When summarizing these studies using discrete treatment
30 levels, the lowest O3 concentration that significantly reduced yield was determined from
31 analyses done by the original authors. Often, the lowest concentration used in the study was the
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Table 9-13. Summary of Ozone Exposure Indices Calculated for 3- or 5-Month
Growing Seasons from 1982 to 1991a
3-month growing season (June-August)
No of
Year Sites"
1982 99
1983 102
1984 104
1985 117
1986 123
1987 121
1988 139
1989 171
1990 188
1991 199
Among Years
HDM2C
ppm
Mean CVd
0.114 23.7%
0.125 24.9%
0.117 24.6%
0.117 24.6%
0.115 21.8%
0.119 22.9%
0.129 21.3%
0.105 23.1%
0.105 21.6%
0.106 22.0%
0.113 11.1%
M7
ppm
Mean
0
0
0
0
0
0
0
0
0
0
0
CV
18.7%
21.9%
18.2%
17.1%
19.1%
17.6%
17.8%
17.5%
18.3%
18.4%
10.0%
SUMOO
ppm-h
Mean
82.9
86.1
84.1
84.6
85.3
86.9
97.6
86.4
85.7
87.7
87.0
CV
19.1%
22.1%
19.9%
18.0%
18.0%
17.3%
19.6%
19.9%
21.0%
21.3%
9.9%
SUM06
ppm-h
Mean CV
26.8 68.8%
34.5 58.1%
27.7 58.4%
27.4 59.6%
27.7 65.0%
31.2 56.4%
45.2 46.8%
24.8 78.7%
25.8 76.2%
28.3 74.2%
29.5 42.1%
SIGMOID
ppm-h
Mean
26.3
33.0
27.4
27.4
27.7
30.4
42.9
25.8
26.6
28.9
29.4
CV
56.7%
52.3%
47.9%
47.6%
51.8%
46.8%
42.4%
59.4%
59.2%
59.5%
31.0%
5-month growing season (May-September)
No. of
Year Sites
1982 88
1983 87
1984 95
1985 114
1986 118
1987 116
1988 134
1989 158
1990 172
1991 190
Among Years
M7
ppm
Mean
0.048
0.051
0.048
0.048
0.048
0.050
0.054
0.047
0.049
0.050
0.049
SUMOO
ppm-h
cv
20.6%
22.1%
18.0%
18.4%
20.3%
20.3%
18.7%
18.6%
19.8%
19.8%
9.8%
Mean
122.9
129.6
126.2
124.5
123.3
128.7
141.7
127.8
129.4
130.6
129.0
CV
22.3%
24.4%
19.1%
19.4%
21.4%
20.4%
22.0%
22.5%
22.7%
23.6%
9.9%
SUM06
ppm-h
Mean
37.3
44.4
36.7
36.2
34.9
42.2
58.0
32.7
34.6
36.8
38.7
CV
70.9%
61.9%
60.8%
63.8%
70.7%
62.0%
50.5%
87.8%
82.7%
80.7%
42.5%
SIGMOID
ppm-h
Mean
37.1
43.8
37.6
37.0
35.6
41.8
55.6
35.2
37.0
38.8
39.6
CV
57.8%
52.7%
46.9%
50.3%
55.7%
50.3%
45.0%
64.1%
62.1%
62.9%
29.8%
"Updated and additional years from data given in Table III of Tingey et al. (1991), where the spatial and temporal variation
in ambient O3 exposures is expressed in terms of several exposure indices.
b Indicates the number of separate monitoring sites included in the analysis; fewer sites had 5 months of available data than
had 3 months of available data.
c The 2HDM index is calculated for sites with at least 3 months of available data. SUMOO, SUM06, M7, SIGMOID, and
2HDM are the cumulative sum above 0.0 ppm, the cumulative sum above 0.06 ppm, the 7-h seasonal mean, the sigmoid
weighted summed concentration, and the second highest daily maximum 1-h concentration, respectively.
d C V = coefficient of variation.
Source: Table 5-30 from U.S. Environmental Protection Agency (1996) based on Tingey etal. (1991).
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1 lowest concentration reported to reduce yield; hence, it was not always possible to estimate a no-
2 effect exposure concentration. In general, the data indicated that 100 ppb O3 (frequently the
3 lowest concentration used in the studies) for a few hours per day for several days to several
4 weeks usually caused significant yield reductions of 10 to 50%.
5 By the time the 1986 document was prepared, much new information concerning the
6 effects of O3 on the yield of crop plants had become available through EPA's NCLAN research
7 program and through research funded by other agencies. The NCLAN project was initiated by
8 EPA in 1980 primarily to improve estimates of yield loss under field conditions and to estimate
9 the magnitude of crop losses caused by O3 throughout the United States (Heck et al., 1982,
10 1991). The cultural conditions used in the NCLAN studies approximated typical agronomic
11 practices. The primary objectives were:
12 (1) to define relationships between yields of major agricultural crops and O3 exposure
as required to provide data necessary for economic assessments and development
ofO3NAAQS;
13 (2) to assess the national economic consequences resulting from O3 exposure of major
agricultural crops; and
14 (3) to advance understanding of cause-and-effect relationships that determine crop
responses to pollutant exposures.
15 Using NCLAN data, the O3 concentrations predicted to cause 10 or 30% yield loss were
16 estimated using linear or Weibull response functions. The data in Table 9-14 are from the 1996
17 document and were based on yield-response functions for 38 species or cultivars developed from
18 studies using OTCs of the type developed by Heagle et al. (1973) (see Section 9.2). Composite
19 exposure-response functions for both crops and tree seedlings as a function of O3 exposure
20 expressed as SUM06 are shown in Figure 9-18. Review of these data indicate that 10% yield
21 reductions could be predicted for more than 50% of experimental cases when: (1) 12-h SUM06
22 values exceeded 24.4 ppm-h, (2) SIGMOID values exceeded 21.5 ppm-h, or (3) 7-h seasonal
23 mean concentrations were 50 ppb. The SIGMOID index is very similar to the W126 index (see
24 Section 9.5 for further information about O3 indices). Much lower values are required for each
25 index to protect 75% of experimental cases (Table 9-15). Grain crops were generally found to
26 be less sensitive than other crops. The data summarized in the 1996 criteria document also
27 indicated that the variation in sensitivity within species may be as great as differences between
28 species.
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Table 9-14. Ozone Exposure Levels (Using Various Indices) Estimated To Cause at
Least 10% Crop Loss in 50 and 75% of Experimental Cases"
50th PERCENTILE"
NCLAN Data (n = 49; wet and dry)d
NCLAN Data (n = 39; wet only)
NCLAN Data (n = 54; wet and dry)6
NCLAN Data (n = 42; wet only)6
NCLAN Data (n = 10; wet)
NCLAN Data (n = 10; dry)
Cotton Data (n = 5)
Soybean Data (n = 13)
Wheat Data (n= 6)
Cotton Data (n = 5)e
Soybean Data (n=15)e
Wheat Data (n = 7)e
SUM06
24.4
22.3
26.4
23.4
25.9
45.7
23.6
26.2
21.3
30.0
23.9
25.9
SEC
3.4
1.0
3.2
3.1
4.5
23.3
2.3
5.4
15.2
12.7
6.5
10.5
SIGMOID
21.5
19.4
23.5
22.9
23.4
40.6
19.3
22.6
19.3
27.2
22.0
21.4
SE
2.0
2.3
2.4
4.7
3.2
0.1
2.3
3.6
12.7
12.8
8.0
9.4
M7
0.049
0.046
NA
NA
0.041
0.059
0.041
0.044
0.061
NA
NA
NA
SE
0.003
0.003
NA
NA
0.001
0.014
0.001
0.005
0.018
NA
NA
NA
2HDM
0.094
0.090
0.099
0.089
0.110
0.119
0.066
0.085
0.098
0.075
0.088
0.097
SE
0.006
0.010
0.011
0.008
0.042
0.017
0.032
0.013
0.059
0.012
0.008
0.028
75th PERCENTILE"
NCLAN Data (n = 49; wet and dry)
NCLAN Data (n = 39; wet only)
NCLAN Data (n = 54; wet and dry)6
NCLAN Data (n = 42; wet only)6
NCLAN Data (n = 10; wet)
NCLAN Data (n = 10; dry)
Cotton Data (n = 5)
Soybean Data (n = 13)
Wheat Data (n = 6)
Cotton Data (n = 5)e
Soybean Data (n=15)e
Wheat Data (n = 7)e
14.2
14.3
16.5
17.2
16.4
24.0
21.8
14.2
11.7
21.1
15.3
5.1
4.2
2.7
4.3
3.0
3.7
0.8
5.0
0.1
2.5
6.0
4.1
2.6
11.9
12.6
14.5
14.7
13.7
22.3
17.5
12.4
10.9
16.7
13.4
8.5
5.6
2.3
3.2
2.4
3.2
0.1
2.8
0.1
2.4
5.7
4.1
3.4
0.040
0.039
NA
NA
0.040
0.053
0.041
0.041
0.054
NA
NA
NA
0.007
0.005
NA
NA
0.001
0.022
0.001
0.006
0.032
NA
NA
NA
0.051
0.056
0.073
0.070
0.080
0.093
0.065
0.069
0.062
0.070
0.078
0.054
0.010
0.006
0.006
0.006
0.032
0.003
0.014
0.004
0.035
0.034
0.007
0.027
aSee Appendix A for abbreviations and acronyms.
bThe numbers in parentheses are the number of cases used in deriving the various exposure levels.
°Standard error (SE).
dNCLAN data refers to studies conducted as part of the NCLAN project. Wet and dry refer to watering regimes
used in the studies, wet being well-watered, and dry meaning some level of drought stress was imposed.
e24-h exposure statistics reported in Lee et al. (1994b). Relative yield loss for 2HDM is relative to yield at 40 ppb
rather than 0 ppb as was used in Tingey et al. (1991).
Source: U.S. Environmental Protection Agency (1996), (modified from Tingey et al. (1991).
January 2005
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o
'•B
100%
90%
80%
70%
60%
50%
40%
30%
20%
10%
0%
a. Crops
75th Percentile
50th Percentile
25th Percentile
\
10
I
20
I
30
I
40
I
50
60
24-h SUM06 (ppm-h)
^_^
«"
o
w
03
E
o
5
•a
•2
1
£
100%-q
90% -1
80% -|
70% -I
60% -I
50% -1
40% -1
30% -|
20% -i
10%-|
n% -
b. Tree Seedlings
75th Percentile
50th Percentile
~~ 25th Percentile
\
10
\
20
\
30
\
40
50 60
24-h SUM06 (ppm-h) (adjusted to 92 days)
Figure 9-18. Distribution of biomass loss predictions from Weibull and linear exposure-
response models that relate biomass to O3 exposure. Exposure is
characterized with the 24-h SUM06 statistic using data from (a) 31 crop
studies from National Crop Loss Assessment Network (NCLAN) and
(b) 26 tree seedling studies conducted at U.S. Environmental Protection
Agencys' Environmental Research Laboratory in Corvallis, OR; Smoky
Mountains National Park, TN; Houghton, Michigan; and Delaware, Ohio.
Separate regressions were calculated for studies with multiple harvests or
cultivars, resulting in a total of 54 individual equations from the 31 NCLAN
studies and 56 equations from the 26 seedling studies. Each equation was
used to calculate the predicted relative yield or biomass loss at 10, 20, 30, 40,
50, and 60 ppm-h, and the distributions of the resulting loss were plotted.
The solid line is the calculated Weibull fit at the 50th percentile.
Source: U.S. Environmental Protection Agency (1996); Hogsett et al. (1995).
January 2005
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1 The chemical protectant, ethylene diurea (EDU), was also used to provide estimates of
2 yield loss. The impact of O3 on yield was determined by comparing the yield data from plots
3 treated with EDU versus untreated plots. Studies indicated that yields were reduced by 18 to
4 41% when daytime ambient O3 concentrations exceeded 80 ppb for 5 to 18 days over the
5 growing season. For this approach to be credible, the effects of EDU itself on a particular
6 species must be preestablished under conditions without O3 exposure (Kostka-Rick and
7 Manning, 1992).
8 The 1996 criteria document reviewed several experiments demonstrating that the seedlings
9 of some tree species such as poplars (Populus) and black cherry are as sensitive to O3 as are
10 annual plants, in spite of the fact that trees are longer-lived and generally have lower rates of gas
11 exchange, and, therefore, a lower uptake of O3. The 1996 document also reviewed data showing
12 that O3 exposures that occur at present in the United States are sufficient to affect the growth of a
13 number of trees species. For example, exposure-response functions for 51 cases of tree seedling
14 responses to O3, including 11 species representing deciduous and evergreen growth habits,
15 suggest that a SUM06 exposure for 5 months of 31.5 ppm-h would protect hardwoods from a
16 10% growth loss in 50% of the cases studied (Table 9-15). Similarly, a SUM06 exposure of
17 42.6 ppm-h should provide the same level of protection for evergreen seedlings. However, these
18 results do not take into the account the possibility of effects on growth in subsequent years.
19 Because multiple-year exposures may cause a cumulative effect on the growth of some trees
20 (Simini et al., 1992; Temple et al., 1992), it is likely that a number of species are currently being
21 affected even at ambient exposures (Table 9-14).
22 In 1986, the EPA (U.S. Environmental Protection Agency, 1986) established that 7-h per
23 day growing season mean exposures to O3 concentrations above 50 ppb were likely to cause
24 measurable yield loss in agricultural crops. At that time, few conclusions could be drawn about
25 the response of deciduous or evergreen trees or shrubs, due to the lack of information about
26 response of such plants to season-long exposures to O3 concentrations of 40 to 60 ppb and above.
27 However, the 1978 and 1986 criteria documents (U.S. Environmental Protection Agency, 1978,
28 1986) indicated that the limiting value for foliar symptoms on trees and shrubs was 60 to 100
29 ppb for 4 h. From 1986 to 1996, extensive research was conducted, establishing the sensitivity
30 of many tree species. Based on research published since the 1986 criteria document,, a number
31 of conclusions were drawn in 1996 AQCD (U.S. Environmental Protection Agency, 1996):
January 2005 9-189 DRAFT-DO NOT QUOTE OR CITE
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Table 9-15. SUM06 Levels Associated with 10 and 20% Total Biomass Loss for
50 and 75% of the Seedling Studies
(The SUM06 value is adjusted to an exposure length of 92 days per year.)a
Weibull Equations (all 51 seedling studies):
50th Percentile PRYL1 = 1 - exp(-[SUM06/176.342]** 1.34962)
75th Percentile PRYL = 1 - exp(-[SUM06/104.281]** 1.46719)
Weibull Equations (27fast-growing seedling studies):
50th Percentile PRYL = 1 - exp(-[SUM06/150.636]** 1.43220)
75th Percentile PRYL = 1 - exp(-[SUM06/89.983]** 1.49261)
Weibull Equations (24 slow to moderate growing seedling studies):
50th Percentile PRYL = 1 - exp(-[SUM06/190.900]** 1.49986)
75th Percentile PRYL = 1 - exp(-[SUM06/172.443]**l.14634)
Weibull Equations (28 deciduous seedling studies):
50th Percentile PRYL = 1 - exp(-[SUM06/142.709]** 1.48845)
75th Percentile PRYL = 1 - exp(-[SUM06/87.724]** 1.53324)
Weibull Equations (23 evergreen seedling studies):
50th Percentile PRYL = 1 - exp(-[SUM06/262.911]** 1.23673)
75th Percentile PRYL = 1 - exp(-[SUM06/201.372]** 1.01470)
Levels Associated with Prevention of a 10 and 20% Total Biomass Loss
for 50 and 75% of the Seedlings
All 51 Seedling Cases
Percent of Seedlings
50% 75%
Relative 10% 33.3 22.5
Biomass Loss 20% 58.0 37.5
27 Fast-Growing Seedling Cases
Percent of Seedlings
Relative
Biomass Loss
10%
20%
50%
31.3
52.9
75%
19.4
32.4
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Table 9-15 (cont'd). SUM06 Levels Associated with 10 and 20% Total Biomass Loss for
50 and 75% of the Seedling Studies
(The SUM06 value is adjusted to an exposure length of 92 days per year.)a
24 Slow-to-Moderate-Growth Seedling Cases
Percent of Seedlings
50% 75%
Relative 70% 42.6 24.2
Biomass Loss 20% 70.2 46.6
28 Deciduous Seedling Cases
Percent of Seedlings
50% 75%
Relative 70% 31.5 20.2
Biomass Loss 20% 52.1 33
23 Evergreen Seedling Cases
Percent of Seedlings
50% 75%
Relative 70% 42.6 21.9
Biomass Loss 20% 78.2 45.9
aSee Appendix [XXX] for abbreviations and acronyms.
bPRYL = predicted relative yield (biomass) loss
Source: U.S. Environmental Protection Agency (1996), based onHogsett et al. (1995).
(1) An analysis of 10 years of monitoring data from more than 80 to almost 200 non-
urban sites in the United States established ambient 7-h growing season average
concentrations of O3 for 3 or 5 months of 51 to 60 ppb and 47 to 54 ppb,
respectively. The SUM06 exposures ranged (a) from 24.8 to 45.2 ppm-h for
3 months and (b) from 32.7 to 58.0 ppm-h for 5 months (Tingey et al. (1991),
Table 9-14).
(2) The results of OTC studies that compared yields at ambient O3 exposures with
those in filtered air and retrospective analyses of crop data (Table 9-14) established
that ambient O3 concentrations were sufficient to reduce the yield of major crops in
the United States. Research results since 1978 did not invalidate EPA conclusions
(U.S. Environmental Protection Agency, 1978, 1986) that foliar symptoms due to
January 2005 9-191 DRAFT-DO NOT QUOTE OR CITE
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O3 exposures reduce the market value of certain crops and ornamentals where
leaves are the product (such as spinach, petunia, geranium, and poinsettia) and that
such damage occurs at ambient O3 concentrations observed in the United States.
3 (3) A 3-month SUM06 exposure of 24.4 ppm-h, corresponding to a 7-h mean of 49 ppb
and a 2HDM of 94 ppb O3 may prevent a 10% loss in 50% of the 49 experimental
cases analyzed by Tingey et al. (1991). A 12-h growing season mean of 0.045 ppb
should restrict yield losses to 10% in major crop species (Lesser et al., 1990).
4 (4) Depending on duration, concentrations of O3 and SUM06 exposures currently in
the United States are sufficient to affect the growth of a number of tree species.
Given the fact that multiple-year exposures may cause a cumulative effect on the
growth of some trees (Simini et al., 1992; Temple et al., 1992), it is likely that a
number of species currently are being impacted, even at ambient O3 exposures
(Table 9-14).
5 (5) Exposure-response functions for 51 cases of seedling response to O3 (Hogsett et al.,
1995), including 11 species representing deciduous and evergreen growth habits,
suggest that a SUM06 exposure for 5 months of 31.5 ppm-h would protect
hardwoods from a 10% growth loss in 50% of the cases studied. A SUM06
exposure of 42.6 ppm-h should provide the same level of protection for evergreen
seedlings. Note that these conclusions do not take into the account the possibility
of effects on growth in subsequent years, an important consideration in the case of
long-lived species.
6 (6) Studies of the response of trees to O3 have established that, in some cases (for
instance, poplars and black cherry), trees are as sensitive to O3 as are annual plants,
in spite of the fact that trees are longer-lived and generally have lower gas
exchange rates, and, therefore, lower O3 uptake.
7 (7) Use of the chemical protectant, EDU, is of value in estimating O3-related losses in
crop yield and tree growth, provided that care is exercised in establishing
appropriate EDU dosages to protect the plants without affecting growth.
8 The major question to be addressed in the remainder of this section is whether new
9 information supports or alters the 1996 criteria document conclusions summarized above.
10 In particular, this section evaluates whether the response of plants to experimental treatments at
11 or near O3 concentrations characteristic of ambient levels in many areas of the United States
12 (Table 9-14) can be compared to a control or reduced O3 treatment to establish a potential
13 adverse effect. Before evaluating new information from the literature on O3 effects on
14 vegetation, O3 exposure indices used in O3 studies and trends in O3 exposure patterns during the
15 past two decades are briefly reviewed.
16
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1 9.6.3 Ozone Indices and Ambient Exposure
2 As recognized in both the 1986 and the 1996 criteria documents, the characterization and
3 representation of the exposure of vegetation to O3 is problematic, because the specific aspects of
4 pollutant exposure that cause injury or damage are difficult to quantify. This issue is addressed
5 in Section 9.5, and only a few points will be discussed here in order to provide a context for
6 interpreting data on exposure-response relationships. The most important effects of O3 on
7 vegetation occur due to uptake of O3 through stomata, with subsequent oxidative injury that
8 appears to be rather nonspecific (Section 9.3). As has been discussed by numerous authors
9 during the last three decades, from a toxicological and physiological view, it is much more
10 realistic to relate effects to internal (absorbed) O3 dose rather than to exposure near the leaf or
11 canopy (Fowler and Cape, 1982; Fuhrer et al., 1992; Griinhage et al., 1993, 1999; Legge et al.,
12 1995; Massman et al., 2000; Musselman and Massman, 1999; Pleijel et al., 1995; Runeckles,
13 1974; Taylor, Jr. et al., 1982; Tingey and Taylor, Jr., 1982) (see also Section 9.5). Theoretically,
14 flux estimates should improve the assessment of O3 effects, but despite recent attention to this
15 topic, particularly in Europe, it remains difficult to estimate flux in the field outside of
16 experimental sites where continuous measurements of wind speed and other environmental
17 conditions are made. This topic is discussed further below in Section 9.6.4.5.
18 No simple exposure index can accurately represent all of the numerous factors operating at
19 different timescales that affect O3 flux into plants and subsequent plant response (Section 9.5).
20 Indices of peaks, such as the 2FIDM are not well suited for discerning exposure-response
21 relationships, because they do not capture the effects of lower O3 concentrations nor the
22 cumulative effects of O3 on vegetation (Heck and Cowling, 1997; U.S. Environmental Protection
23 Agency, 1996). For this reason, peak indices have not been used in recent decades to develop
24 exposure-response relationships for vegetation. Fortunately, other simple indices have shown
25 substantial correlation with responses such as crop yield under experimental conditions. During
26 the 1980s, the most commonly used indices for expressing O3 exposure were 7-, 8-, or 12-h
27 daytime average values over the duration of O3 exposure, which was often 3 months or
28 somewhat less for experimental studies with crops. These indices perform reasonably well for
29 interpreting experimental data on the response of vegetation to ozone, particularly for individual
30 experiments, although they do not explain all of the variation among experiments in
31 retrospective analysis of multiple experiments (Lesser et al., 1990).
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1 Since the 1980s, cumulative indices such as the SUM06, AOT40, or W126 that
2 preferentially weight higher concentrations have been used in conjunction with mean indices for
3 developing exposure-response relationships (Tables 9-15 and 9-16, and Figure 9-1). Such
4 indices are often more suitable than mean values because they are cumulative and because they
5 preferentially weight higher concentrations. Thus, these indices generally provide somewhat
6 better fits to experimental data than do mean indices, especially in retrospective analyses of
7 multiple experiments on multiple species (Lee et al., 1994a, 1994b; Lee and Hogsett, 1999;
8 Tingey et al., 1991). Unfortunately, no single index has been used consistently even in the
9 recent literature, making it difficult to compare results among and between experiments and with
10 ambient exposure data. However, Tables 9-14 and 9-17 provide summaries of ambient exposure
11 data for several indices that can be compared to the experimental results reviewed in the
12 remainder of this section. Of the cumulative indices that preferentially weight higher
13 concentrations, the SUM06 index has been used most commonly in the U.S. literature, and it was
14 selected in a meeting of scientific experts on O3 effects on vegetation as suitable for a secondary
15 standard to protect vegetation (Heck and Cowling, 1997). However, it should be noted that the
16 W126 index has been selected for use in protecting vegetation in Class 1 areas (Federal Land
17 Managers' Air Quality Related Values Workgroup (FLAG), 2000). Even in recent studies, O3
18 data are often presented using only a seasonal mean index value, and so mean values are
19 frequently presented in this section. Such reporting of mean indices should not be interpreted as
20 a preference for them, but rather as a limitation in the data reported in the literature. Additional
21 information about O3 exposure for individual experiments, including the number and type of O3
22 treatments (addition of a constant concentration of O3 or an amount proportional to ambient
23 levels), and duration, are reported in Tables 9-16 through 9-19.
24 Since the 1996 document, the use of the AOT40 index has become quite common in
25 Europe for identifying and mapping areas of exceedance, but it has not been used much in the
26 United States. Thus, studies reporting O3 exposure only as AOT40 values are presented in tables
27 summarizing effects on annual, herbaceous perennial, and woody vegetation. However, such
28 studies are not as commonly cited in the text of this section because AOT40 summary data on O3
29 exposures in the United States are rarely available. This lack makes it difficult to compare
30 experimentally derived exposure-response data expressed as AOT40 to ambient U.S. O3
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3
Table 9-16. Summary of Selected Studies of Ozone Effects on Annual Species
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Species
Bean, cv.
Pros
Bean, cv. Lit
Corn
Cotton, cv.
Deltapine
Cotton, cv.
Deltapine
Oat, cv. Vital
Potato2
Rape, oilseed
Rice, cvs.
Koshi-hikari,
Nippon-bare
Soybean
Soybean, cv.
Essex
Facility
OTC
OTC
OTC
OTC
OTC
OTC
OTC
Open Air
OTC
OTC
OTC
Location
The Netherlands
Germany
Beltsville, MD
Raleigh, NC
Raleigh, NC
Ostad, Sweden
6 sites N. Europe
Northumberland,
UK
Japan
Beltsville, MD
Raleigh, NC
O3 Concentration
(Units are ppb unless
otherwise specified)1
CFtoCF75:
9-hmean = 3-70,
AOT40 = Oto 17.7ppm«h
CF,NF, CF-lx,CF-2x:
mean=l, 14,15,32
CF, +40: 7-h mean = 20, 70
CF, 1.5x:
12-hmean = 21,71
CF,NF, 1.5x,
12-hmean = 24, 51, 78
CF,NF: 7-h mean =12, 27
AOT40 = 6-27 ppm«h
AA, +O3: 7-h mean for
17 days Aug. -Sept = 30, 77,
for 32 days in
May-June = 31, 80
CF, lx, 1.5x,2x,2.75x:7-h
mean= 13.5-93.4
CF, +40: 7-h mean = 25, 72
CF, 1.5 x: 12-h mean for
3 years = 23, 82
Duration
62 days
3 months
1 years
1 years
1 years
1 years
2 years (1 year
at 2 sites)
17 days in fall,
overwinter,
32 days in spring
3 years
2 years
3 years
Variable
g=Green pod
yield
Pod yield
Grain yield
Seed-cotton
weight
Seed-cotton
weight
Grain yield
Tuber yield
Seed yield
Grain yield
Seed yield
Seed yield
Response
(Decrease from
lowest, %)
29 at 9-h mean = 44
(AOT40 = 3.6ppm«h)
56 (CF, 2x)
13
22
21,49(NF, 1.5x)
+2 (n.s.)
4% average for
all experiments
14
3 to 10 at 40 ppb
25
41
Reference
Tonneijck and
Van Dijk (1998)
Brunschon-Harti
etal. (1995)
Mulchi et al.
(1995)Rudorff
etal. (1996c)
Heagle et al.
(1999)
Heagle et al.
(1999)
Pleijel et al.
(1994a)
Craigon et al.
(2002)
Ollerenshaw
etal. (1999)
Kobayashi et al.
(1995)
Mulchi et al.
(1995)
Fiscus et al.
(1997)
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Table 9-16 (cont'd). Summary of Selected Studies of Ozone Effects on
Species
Soybean, cvs.
Forrest, Essex
Soybean, cv.
Essex
Soybean, cv.
Essex
Soybean, cv.
Holladay
Soybean, cv.
NK-6955
Soybean,
3 cvs.
Soybean,
3 cvs.
Soybean, cv.
Essex
Soybean, cvs,
Essex, Forrest
Timothy
Watermelon
Facility
OTC
OTC
OTC
OTC
OTC
OTC
OTC
OTC
OTC
OTC
OTC
Location
Maryland
Raleigh, NC
Raleigh, NC
Raleigh, NC
Raleigh, NC
Raleigh, NC
Raleigh, NC
Raleigh, NC
Beltsville, MD
Sweden
Spain
O3 Concentration
(Units are ppb unless
otherwise specified)1
CF, +40: 7-h mean = 24 and
24, 63 and 62 for each year
CF,NF, 1.5x:
12-h mean = 20, 50,79
CF,NF, 1.5x:
12-h mean =18, 42,69
CF,NF, 1.5x:
12-h mean = 18,42,69
CF,NF, 1.5x:
12-h mean = 18,42,69
CF,NF, 1.5x:
12-h mean =14, 36,64
CF,NF, 1.5x: 12-h mean =
24, 49, 83
CF,NF, 1.5x:
12-h mean = 20, 50, 79
CF,NF+: 7-h mean = 24, 58
AOT40 = 10, 20, 340;
12-h mean = 20, 152
CF(O3 = 0),NFml988
AOT40 = 5.96 ppm«h,
SUM06 = 0.29 ppm«h, in
1989 AOT40 =18.92 ppm«h,
SUM06 = 4.95ppm«h
Duration Variable
2 years Seed yield
1 year Seed yield
1 year Seed yield
1 year Seed yield
1 year Seed yield
3 months Seed yield
4 months Seed yield
4 months Seed yield
134 days Seed yield
1 year Biomass
2 expts of 1 year Fruit yield
Annual Species
Response
(Decrease from
lowest, %)
10, 32 (2 cvs.)
16, 37 (NF, 1.5x)
15, 40 (NF, 1.5x)
22, 36 (NF, 1.5x)
+46, +4(NF, 1.5x)
At ambient = +14, 11,
16 for 3 cvs.
At ambient = 17, 13,
18 (3 cvs.)
ll,22(amb., 1.5x)
Essex = +11 (ns),
Forrest = 21
58
19, 3 9 (2 expts)
Reference
Chernikova
et al. (2000)
Heagle et al.
(1998)
Heagle et al.
(1998)
Heagle et al.
(1998)
Heagle et al.
(1998)
Miller et al.
(1994)
Miller et al.
(1994)
Miller et al.
(1998)
Robinson and
Britz (2000)
Danielsson et al.
(1999)
Gimeno et al.
(1999)
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Table 9-16 (cont'd). Summary of Selected Studies of Ozone Effects on Annual Species
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Species Facility
Wheat1, cv. OTC
Minaret
Wheat1 OTC
Wheat, cv. OTC
promessa
Wheat, cv. OTC
promessa
Wheat, cv. OTC
promessa
Wheat, cvs. OTC
Massey,
Saluda
Wheat, cv. OTC
Turbo
Wheat, cv. OTC
Turbo
Wheat, cv. OTC
Turbo
Location
8 sites in
N Europe
Sweden
SE Ireland
SE Ireland
SE Ireland
Beltsville, MD
Germany
Germany
Germany
O3 Concentration
(Units are ppb unless
otherwise specified)1
12-h mean (SD) low = 26.3
(12.2), 12-h mean (SD)-high
= 51.4 (18.3)
AOT40 mean (SD)
low = 6.18(8.54)ppm«h,
AOT40 mean (SD)
high = 28.23 (23.05) ppm«h
AOT40 0 to 15 ppm«h
CF, +50:
12-h total = 5.6, 32.6 ppm«h
CF, +25:
12-h total = 6.2, 33.4 ppm«h
CF, +25, +50:
12-h total = 6.7, 34,
34 ppm«h
CF, +40:
7-h mean =19, 20 and 61,
65 (2 years)
8-h mean = 5. 9, 61.2, 92.5
8-h mean = 4. 7, 86.4
7-h mean = 5, 41,73
Duration
13 studies of
1 year each
7 years
3 h/day,
5 d/week,
7 weeks
6 h/day,
5 d/week,
7 weeks
+25 = 6 h/day,
5 d/week,
+50 = 3 h/day,
5 d/week,
both 7 weeks
2 years
1 year
1 year
1 year
Variable
Grain yield
Grain yield
Grain yield
Grain yield
Grain yield
Grain yield
Grain yield
Grain yield
Grain yield
Response (Decrease
from lowest, %)
13 (n.s.)
23 at AOT40 =
15 ppm«h
53
+17
Amb + 25 = 3 (n.s.);
Amb + 50 = 17
20
14,40
(mid, high O3)
20
35
Reference
Bender et al.
(1999)Hertstein
etal. (1999)
Danielsson et al.
(2003)
Finnan et al.
(1996a)
Finnan et al.
(1996a)
Finnan et al.
(1996a)
Mulchi et al.
(1995);Rudorff
etal. (1996b)
Bender et al.
(1994)
Bender et al.
(1994)
Fangmeier et al.
(1994)
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Table 9-16 (cont'd). Summary of Selected Studies of Ozone Effects on Annual Species
O3 Concentration
(Units are ppb unless Response (Decrease
Species Facility Location otherwise specified)1 Duration Variable from lowest, %)
Wheat, OTC Raleigh, NC 1 2-h mean = 27, 47, 90 2 months Grain yield 5 (n.s.)
winter, 8 cvs.
Wheat, OTC Raleigh, NC 1 2-h mean = 22, 38, 74 2 months Grain yield 16 (n.s.)
winter, 8 cvs
Wheat, cv. OTC Finland 1992: 12-hmean = 14, 30, 2 years Grain yield At highest O3 =
Drabant 61; AOT 40 = 16.3, 34.8, 13 each year
54.6ppm«h. 1993:
12-hmean =9, 21,45;
AOT40 = 10.2, 24.8,
40.6 ppm«h
Wheat, cv. Open Air Northumberland, AOT40 for Mar to 1 year, Grain yield 1 3
Riband UK Aug 93 = 3.5, 6.2 ppm«h overwinter
Values for ambient or NF treatments are indicated in bold.
2
Bold indicates that multiple experiments (more than just 2 years at a single site) were included in the analysis.
Reference
Heagle et al.
(2000)
Heagle et al.
(2000)
Ojanpera et al.
(1998)
Ollerenshaw and
Lyons (1999)
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Table 9-17. Summary of Selected Studies of the
Species
Alfalfa, cvs.
Apica, Team
Bahia grass
Bent grass
(Capillaris sp.)
Blackberry
Clover, white
Clover, white
Clover, white
Clover, white
and red
Clover, white,
cv. Menna
Facility Location
OTC Quebec,
Canada
OTC Auburn, AL
OTC United
Kingdom
Large Alabama
OTC
Ambient MA, OR, NC,
air CA (2 sites)
andVA
Ambient 14 European
air sites
OTC United
Kingdom
OTC Switzerland
OTC Italy
O3 Concentration
(Units are ppb unless
otherwise specified)1
12-hmean: 1991 = 6,39,49,110;
1992 = 0,34,42,94
12-hmean = 22, 45, 91
AOT40 = 0.8-15.0ppm-h
1994: AOT40 = 2-1 12 ppm-h,
SUM06= l-162ppm-h,
1995: AOT40 = 3-83 ppm-h,
SUM06 = 0-132ppm-h
SUM06 for 6-h/day = 10.2-39.4
ppm-h, AOT40 for 12-h/day =
0.6-50.1 ppm-h
AOT40 for 28 d = 0-12 ppm-h
AOT40 = 0.8-15.0ppm-h
CF, NF, NF+. NF++:
12-hmean = 21, 39, 47, 65
CF,NF: AOT40 = 0.1;
8.9 ppm-h, 7-h mean = 24, 53
Effects of Ozone on Perennial Herbaceous Plants
Duration
3 months in
each of
2 years
24 weeks
8 h/day for
3 months
7 months
in 1994,
6 months
in 1995
2 growing
seasons
3 growing
seasons
8 h/day for
3 months
3.5 months/
year for
2 years
2 months
Variable
Biomass
Biomass at ambient O3 for
1 st, 2nd cutting of early
and late season plantings
Biomass, in competition
with 3 other spp. Total
biomass in uncut pots,
aboveground biomass in
cut pots (cut every
14 days).
Percent canopy cover
(grown in old field
community), biomass ripe
fruit number
Biomass ratio
(sensitive/resistant)
Biomass ratio
sensitive/resistant)
Biomass, in competition
3 other spp. Total biomass
in uncut pots, aboveground
biomass in cut pots (cut
every 14 days).
Biomass, in managed
pasture
Biomass
Response (Decrease
from lowest, %)
ForNF: Apica = 31,
21; Team = 14,2
(n.s.)
34,29(n.s.),+6(n.s.),
9 (n.s.)
8 (uncut), +18 (cut)
+124 for cover, n.s.
for biomass, 28% for
ripe fruit number but
sig. chamber effect.
4 at 6-h SUMO6 =
39.4 ppm-h; 12 h
AOT40 = 50.1ppm-h
5atAOT40for
28 days = 0.9-1. 7
ppm-h
18 (uncut), 40 (cut)
24, 26, 52
20
Reference
Renaud et al.
(1997)
Muntifering
et al. ( 2000)
Ashmore and
Ainsworth
(1995)
Barbo et al.
(1998)
Chappelka
(2002)
Heagle and
Stefanski
(2000)
Mills et al.
(2000)
Ashmore and
Ainsworth,
(1995)
Fuhrer et al.
(1994)
Fumagalli
etal. (1997)
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lauie ?-i / ^
Species Facility
Fescue, red OTC
Lespediza, OTC
Sericea
Little bluestem OTC
Phleum alpinum OTC
Rasberry OTC
Speedwell, OTC
Germander
Strawberry OTC
Sumac, winged OTC
Timothy OTC
com u>. CMIIIII
Location
United
Kingdom
Auburn, AL
Auburn, AL
Sweden
Ontario
United
Kingdom
United
Kingdon
Alabama
Sweden
nary 01 aeiecieu aiuuies
O3 Concentration
(Units are ppb unless
otherwise specified)1
AOT40 = 0.8-15.0ppm-h
CF,NF,2x: 12-hmean = 23,
40, 83, SUM06 = 0.2, 9.1, 61.0,
AOT40 = 0.6, 7.0, 39.8
CF,NF,2x: 12-hmean = 23,
40, 83 ppb, SUM06 = 0.2,9.1,
61.0, AOT40 = 0.6, 7.0, 39.8
AOT40 = 0.01, 0.02, 0.34 ppm-h;
12-hmean = 20, 152
lx,+12+24: Ambient < 4
AOT40 = 0.8-15.0ppm-h
8-h mean = 27, 92;
AOT40 for +O3 = 24.59 ppm-h
SUM06 = 0 to 132 ppm-h
CF, NF, CF+:
AOT40 = 0.0, 1.3, 20.3 ppm-h;
12-hmean = 20, 68, 152
01 me i^iiecu
Duration
8 h/day for
3 months
10 weeks
10 weeks
1 year
1 day /week
for 7 weeks
8 h/day for
3 months
69 days
6 months
1 year
> 01 uzoiie on reremii
Variable
Biomass, in competition
with 3 other spp. Total
biomass in uncut pots,
aboveground biomass in
cut pots (cut every
14 days).
Biomass
Biomass
biomass
Fruit yield
Biomass, in competition
with 3 other spp. Total
biomass in uncut pots,
aboveground biomass in
cut pots (cut every 14
days).
Fruit size, yield
Percent canopy cover
(grown in old field
community)
Biomass
ai neruaceous riams
Response (Decrease
from lowest, %) Reference
+ 30 (uncut), Ashmore and
+13 (cut) Ainsworth,
(1995)
n.s. Powell et al.
(2003)
n.s. Powell et al.
(2003)
87 (Danielsson
etal. (1999)
No effect at + 12, Sullivan et al.
52 at +24 in only one ( 1 994)
of two cultivars
14 (uncut), 26 (cut) Ashmore and
Ainsworth
(1995)
Size = 14, Drogoudi and
yield = (n.s.) Ashmore
(2000)
95 (Barbo et al.
(1998)
ns in NF, 58 in CF+ Danielsson
etal. (1999)
1 Values for ambient or NF treatments are indicated in bold.
2 Bold indicates that multiple
experiments (more
than just 2 years at a single site)
were included in the analysis.
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Table 9-18. Summary of Selected Studies of Ozone Effects on Deciduous Trees and Shrubs
Species Age Facility
Ash, OTC
European
Ash, Seedling OTC
European
Aspen Cutting, OTC
Seedling
Aspen Cutting OTC
Aspen Cutting FACE
Aspen Cutting Large OTC
Aspen First year OTC
Beech, OTC
European
Beech, Seedling OTC
European
O3 Concentration
(Units are ppb unless
Location otherwise specified)1 Duration
Hampshire, UK NF, NF+: Mean = 17.7, 3 years for
44.1; AOT40 for 24h = 1.9, day 100 to
59.9 ppm-h day 162
Switzerland 0.5x, 0.85x, lxs 0.5x+30: 5 months
AOT40 = 0.1,3.4, 7.1,
19.7ppm-h
Michigan CF, lx, 2x; 3 months 7-h 98 days
mean for 1990 = 7-69;
for 199 1=22-92
Michigan CF, lx,2x: 98 days
SUMOO= 11,58, 71 ppm-h
Wisconsin ambient, +90 (exposure 3 years
data not reported)
New York lx, 1.7x, 3x: 92 days
SUM06 = 1, 20, 62 ppm-h;
9-h mean = 40, 74, 124
Pennsylvania 8-h mean = 39, 73 11 weeks
Switzerland 0.5x, 0.85x, lx; 0.5x+30: 5 months
AOT40 = 0.1,3.4, 7.1,
19.7 ppm'h
Belgium CF, NF, +30: 23 April -
8-h mean = 5. 29, 33; 30 Sept
AOT40 = 0. 4.06,
8.88 ppm-h
Variable
Growth and
biomass of
organs
Biomass
Total
biomass
Total
biomass
Volume
(d2*h)
Shoot
biomass
Biomass
Biomass
Growth
Response
(Decrease from
lowest, %)
n.s.
26 at lx, 50 at
O.Sx + 30
nsat lx,29at
2x for all clones
in each year
25-38 at 1 x
21 at +90
14, 25 for
2 clones at 1.7x
14-30 for 3 of
6 N treatments
6 at lx, 30 at
O.Sx+30
No effect
Reference
Broadmeadow
and Jackson
(2000)
Landolt et al.
(2000)
Kamosky et al.
(1996)
Dickson et al.
(2001)
Isebrands et al.
(2001)
Yunand
Laurence
(1999a)
Pell et al.
(1995)
Landolt et al.
(2000)
Border et al.
(2000a)
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Table 9-18 (cont'd). Summary of Selected Studies of Ozone Effects on Deciduous Trees
O3 Concentration
(Units are ppb unless
Species Age Facility Location otherwise specified)1
Beech, European Seedling Growth chamber Belgium CF, CF+40, CF+100:
SumO = 0.48, 8.93,25.14
ppm-h; AOT40of
NF+100 = 13.91 ppm-h;
uptake = 159,29657095
molm2
Beech, European 0-3 years OTC Switzerland AOT40 for 24 h/days =
4-73 ppm'h
Beech, Japanese 4 years Growth chamber Japan CF, +60 ppb for 7 h/day
Birch, silver Sapling FACE Finland AOT40 = 1, 15 ppm-h;
7-h mean = 26, 40
Birch, silver Sapling OTC Sweden NF, NF+, NF++, daylight
mean 1997 = 29, 37, 54;
1998 = 25, 42, 71 ppb;
AOT40 1997 = 2.4, 6.9,
35.1; 1998 = 0.6,19.6,
74.7 ppm'h
Birch, Seedling Chamber in Norway AOT40 = 0.1, 2.5, 7.1,
[B. pubescens] glasshouse 7.4, 17.8, 19.8 ppm'h
Cherry, black 2 years OTC Norris, TN CF, lx,2x:
7-h mean = 2 1,50, 97
Cherry, black Seedling OTC GSMNP1 CF, lx, 1.5x, 2x:
SUM06 = 0-40.6 ppm-h,
AOT40 = 0.03-28.3
ppm-h
Duration Variable
7 episodes of Biomass,
5 days diameter
1-3 years Total
biomass
156 days Total
biomass
5 years Biomass
2 years Total
biomass
40 days Biomass
April to August Biomass
76 days Biomass
and Shrubs
Response
(Decrease from
lowest, %)
No effect
20atAOT40
for 24 h = 32
ppm'h
19
34 for root, ns
for stem
Total biomass
nsatNF+, 22 at
NF++; root
biomass 30 at
NF++
Sig. decrease in
rootatAOT40
= 2.5 ppm-h,
shoot at
7.1 ppm'h
No effect
n.s. at lx and
1.5x, 38at2x
Reference
Border et al.
(2001)
Braun and
Fluckiger
(1995)
Yonekura
etal. (2001)
Oksanen
etal. (2001)
Karlsson
et al. (2003)
Mortensen
(1998)
Samuelson
(1994)
Neufeld et al.
(1995)
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Species Age Facility
Cherry, black Seedling OTC
Cherry, black 1 year OTC
Cherry, black Seedling OTC
Cottonwood, Cutting Ambient,
Eastern in buried pots
with irrigation
Grape 3 years OTC
Oak Seedling OTC
Oak, red Seedling OTC
O3 Concentration
(Units are ppb unless
Location otherwise specified)1
GSMNP1 CF, 0.5x, Ixj.Sx^x:
SUM06 = 0-53.7 ppm-h;
AOT40 = 0-40.4 ppm-h
Delaware, OH CF, 0.5, 1, 1.5, 2x:
SUMOO in 1990 = 17-107
ppm-h, in 1991 = 31-197
ppm'h
Pennsylvania CF, 0.75x, 0.97x:
7-h mean = 39 to 46,
SUM06 = 0-10.34 ppm-h
In and within 100 12 h mean = 23-49 ppb
km of New York
City, NY
Austria CF, lx, +30, +50:
(AOT40 = O-50 ppm-h
Hampshire, UK NF, NF+:
Mean = 17.7, 44,
AOT40for24h=1.9,
59.9 ppm-h
Norris, TN SUM06 for 3 years = 0,
29, 326 ppm-h; SUMOO
for 3 years = 147, 255 and
507 ppm'h
Duration
140 days
2 years (in 1990
for 3.5 months,
1991 for
4 months)
3 years for
17 weeks
2 months
each year,
3 10-years
experiments
2 years
(preflowering,
past harvest)
3 years for day
100 to day 162
3 years
Variable
Biomass
Total
biomass
Total
biomass
Total
biomass
Fruit yield
Biomass of
organs
Total
biomass
Response
(Decrease from
lowest, %)
n.s. at lx and
1.5x, 59at2x
no effect at 1 x
and 1.5x, 32 at
2x
6 at 0.75 x, 14 at
0.97x
33% decrease at
38 ppb
compared to
23 ppb
Calculated 10
atAOT40 =
27 ppm'h
30 for total
biomass
n.s.
Reference
Neufeld et al.
(1995)
Rebbeck
(1996)
Kouterick
et al. (2000)
Gregg et al.
(2003)
Soja et al.
(1997)
Broadmeado
wand
Jackson
(2000)
Samuelson
etal. (1996)
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g Table 9-18 (cont'd). Summary of Selected Studies of Ozone Effects on Deciduous Trees
to
o
O Species Age Facility
Oak, red 30 years OTC
Maple, red 2 years OTC
Maple, sugar 1 year OTC
Maple, sugar Seedling Large OTC
VO
to
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Maple, sugar Seedling Large OTC
o
5d Plum, Casselman Sapling Large OTC
M
H
§ Poplar, black Seedling OTC
*
0
o
O3 Concentration
(Units are ppb unless
Location otherwise specified)1
Norris, TN SUM06 for 3 years = 0,
29, 326 ppm-h; SUMOO
for 3 years = 147, 255 and
507 ppm-h
Norris, TN CF, lx,2x:
7-h mean = 2 1,50,
97 ppm-h
Delaware, OH CF, 0.5, 1.5, 2x:
UMOOinl990=17to
107ppm-h, in 1991 =31
to 1 97 ppm'h
Ithaca,NY CF, lx, 1.5x,2x: 3 years
SUMOO = 148 to
591 ppm-h; daytime
mean= 19. 7 to 40. 7
Ithaca,NY lx,1.7x, 3x:
3 years 12-h mean = 38,
69,117
Fresno, CA CF, lx,+O3:
12-h mean =31, 48, 91
Belgium CF, NF, +30:
8-h mean = 5, 29, 33;
AOT40 = 0, 4,8.9 ppm-h
Duration
3 years
April to August
2 years (in 1990
for 3.5 months,
1991 for
4 months)
3 years for 134,
128, 109 days
3 years for 109,
143, 116 days
4 years
23 April -
30 Sept
Variable
Stem
increment
Biomass
Total
biomass
Biomass
Total
biomass
Stem
increment,
fruit yield
Diameter,
height
and Shrubs
Response
(Decrease from
lowest, %)
n.s. despite 50%
reduction in net
photosynthesis
No effect
n.s., but linear
trend
No effect
For 1.7x and
3x: 21, 64 in
low light, 26
and 41 in high
light
Fruit yield 16 at
lx, stem +14 at
+03
29 for diameter
in NF+, no
effect on height
Reference
Samuelson
etal. (1996)
Samuelson
(1994)
Rebbeck
(1996)
Laurence
etal. (1996)
Topa et al.
(2001)
Retzlaffetal.
(1997)
Bortier et al.
(2000b)
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Table 9-18 (cont'd).
Species Age Facility
Poplar, hybrid 0 year FACE
(P. tremuloides
x P. tremula)
Poplar, hybrid Cutting OTC
Yellow-poplar 1 year OTC
Summary of Selected Studies of Ozone Effects on Deciduous Trees and Shrubs
O3 Concentration
(Units are ppb unless
Location otherwise specified)1
Finland AOT40 = 0.07, 1.6
ppm'h; 7-h mean = 30, 38
Michigan CF, CF+100:
12, 48 ppm-h
Delaware, OH CF, 0.5, 1.5, 2x:
SUMOOinl990 = 17to
107ppm-h, in 1991 =31
to 1 97 ppm-h
Duration
2 months
60 days
2 years (in 1990
for 3.5 months,
1991 for
4 months)
Variable
Biomass,
height
Total
biomass
Total
biomass
Response
(Decrease from
lowest, %)
n.s. for biomass,
6 for height
46 for average
of 5 clones
No effect
Reference
Oksanen
etal. (2001)
Dickson et al.
(1998)
Rebbeck
(1996)
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1 Values for ambient or NF treatments are indicated in bold.
2 Bold indicates that multiple experiments (more than just 2 years at a single site) were included in the analysis.
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Fir,
Douglas
Hemlock,
eastern
Pine,
loblolly
Pine,
loblolly
Table 9-19. Summary of Selected Studies of Ozone Effects on Evergreen Trees and Shrubs
Age Facility Location
Seedling Open air British
Columbia
Seedling OTC GSMNP1,
TN
12 weeks OTC Oak Ridge,
TN
1 year OTC Alabama
O3 Concentration
(Units are ppb unless
otherwise specified)1
12 tits:
12-h mean 1988 =18-41;
1989 = 27-66
CFto2x:
SUM06 = 0.2-108.1 ppm«h,
AOT40 = 0.2-63. 9 ppm«h
CFto2x:
24-h summer = 74, 137, 169,
206, 284 ppm«h
1994:
AOT40 = 2-112ppm«h,
Duration
1988 = 92
days; 1989 =
101 days
3 years
3 months
2 years, April
to October
Variable
Second flush
biomass
Biomass
Biomass
Dry weight,
height,
Response
(decrease from
lowest, %)
Calculated 55 at
highest exposure
No effect
14 in lx (avgfor all
families)
n.s.
Reference
Runeckles and
Wright (1996)
Neufeld et al.
(2000)
McLaughlin
etal. (1994)
Barbo et al.
(2002)
VO
SUM06
1995:
diameter
«h,
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Pine, 3 years
loblolly
Pine, Seedling
ponderosa
Pine, 39 to
ponderosa 45 years
Pine, Seedling
ponderosa
SUM06 = 0-132ppm«h
OTC Raleigh, NC Ambient, CF, NF, 1 .5 x .,
2.5x: 12-h mean = 54, 29, 47,
76,98
OTC Corvalhs, For CF 12-h SUM06 =
OR 0 ppm«h; for +03 12-h
SUMO6 = 22, 27, 31ppm«h
for 3 years
Ambient CA 24-h mean for 3 weeks late
gradient July and early August for
1993 and 1994 = 70-90 ppb
OTC CA CF, lx,2x:
24-h mean
approx. 20, 60, 120
5 months
3 years:
16 weeks,
16 weeks,
14 weeks
Ambient
gradient
Height,
diameter,
needle length
Total biomass
Fine and
medium root
growth
Total biomass
No effect on stem
height or diameter,
decrease in needle
length
No effect without
grass, 25 with grass
present
85 at most polluted
site.
n.s.
Anttonen et
(1996)
Andersen et
(2001)
Grulke et al.
(1998)
al.
al.
Takemoto et al.
(1997)
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Table 9-19 (cont'd). Summary of Selected Studies of Ozone Effects on Evergreen Trees and Shrubs
Species
Pine,
Scots
Pine,
Scots
Pine,
Scots
Pine,
Scots
Pine,
Scots
Pine,
Table
Mountain
Pine,
Virginia
Sequoia,
giant
Spruce,
Norway
O3 Concentration
(Units are ppb unless
Age Facility Location otherwise specified)1
OTC Hampshire, NF, NF+:
UK Mean = 17.7, 44.1; 24-h
AOT40 = 1.9, 59.9 ppm«h
3 -6 years Free air Finland Amb, +O3:
AOT40 = 0-1, 2-13 ppm«h
Seedling OTC Switerzland 0.5x, 0.85x, lx, 0.5x+30:
AOT40 = 0.1, 3.4,7.1,
19.7 ppm«h
3 years OTC Finland CF, lx,+O3:
24 h AOT40 for
2 years = 0.5, 6, 73 ppm«h
3 years Free air Finland lx,+O3:
24 h AOT40 for 2 years = 2,
37 ppm«h
Seedling OTC GSMNP2, CFto2x:
TN SUM06 = 0.2-116.4 ppm«h,
AOT40 = 0.2-71. 7 ppm«h
Seedling OTC GSMNP2, CFto2x:
TN SUM06 = 0.1-32.8, 47.9,
56.2 ppm«h;
AOT40 = 0.1-19.3, 27.1, 34.4
ppm«h
125 years Branch California 0.25x, lx; 2x, 3x;
chamber 24-h SUMOO approx. 10, 85,
180, 560ppm«h
4-7 years Open air Finland Amb, +O3:
AOT40 = 01,2-13ppm«h
Duration
3 years for
62 days
3 years
5 months
2 years
(4 months
each)
3 years
(3-4 months
each)
3 years
1 -2 years
(3 expts)
61 days
3 years
Response
(decrease from
Variable lowest, %)
Total biomass 15
Biomass No effect
Biomass 14atlx, 22 at
0.5x+30
Biomass No effect
Root and shoot 32 only for root
biomass biomass in high N
treatment
Biomass Slight decrease in
older needle mass
only
Biomass No effect
Branch growth No effect
Biomass No effect
Reference
Broadmeadow
and Jackson
(2000)
Kainulainen
et al. (2000)
Landolt et al.
(2000)
Utriainen et al.
(2000)
Utriainen and
Holopainen
(2001)
(Neufeld et al.
(2000)
Neufeld et al.
(2000)
Grulke et al.
(1996)
Kainulainen
et al. (2000)
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Table 9-19 (cont'd). Summary of Selected Studies of Ozone Effects on Evergreen Trees and Shrubs
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Species Age Facility Location
Spruce, 3 -7 years OTC Sweden
Norway
Spruce, Seedling OTC Switerzland
Norway
Spruce, 0-3 years OTC Switzerland
Norway
Spruce, Seedling OTC Sweden
Norway
Spruce, Sapling Large Ithaca, NY
red OTC
1 Values for ambient or NF treatments are indicated in
2 Great Smoky Mountains National Park.
O3 Concentration
(Units are ppb unless
otherwise specified)1 Duration
CF, 1.5x; 4 years
12-h mean for 4 years = 12,
44; AOT40 = 2, 23 ppm«h
0.5x,0.85x, lx,0.5x+30: 5 months
AOT40 = 0.1, 3.4,7.1,
19.7 ppm«h
AOT40 for 24 h for 1 to 1-3 years
3 years = 22 to 63 ppm«h
CF, lx, 1.5x; 4 years
AOT40 daylight for 4 years =
1,16, 79ppm«h
CF, lx, 1.5x, 2x: 4 years:
total for 4 years = 21 1 to 98-124
569 ppm«h; days/year
daytime mean = 21-71
bold.
Response
(decrease from
Variable lowest, %) Reference
Total biomass 8 Karlsson et al.
(2002)
Biomass n.s. Landolt et al.
(2000)
Total biomass n.s. Braun and
Fluckiger
(1995)
Stem volume No effect in final Wallin et al.
year (2002)
Biomass No effect Laurence et al.
(1997)
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1 exposures. The development of critical levels in Europe has been based primarily on the AOT40
2 index, so this index is discussed in that context.
3 In addition to peak-weighting, there is also evidence that the timing of exposure during
4 plant growth is important. For example, the greatest effects on grain yield are due to exposure
5 during grain filling, rather than earlier or later in the growing season (Lee et al. 1988; Pleijel
6 et al., 1998; Soja et al., 2000; U.S. Environmental Protection Agency, 1996; Younglove et al.,
7 1994). The importance of respite times was also discussed in the previous criteria documents
8 (U.S. EPA, 1978, 1988, 1996) but remains difficult to quantify (Section 9.5). Even when some
9 of these aspects of O3 exposure can be elucidated, it is difficult to apply this knowledge to
10 developing exposure-response relationships based on data in the scientific literature, because O3
11 exposure is often reported only in the form of a summary index such as a 12- or 24-h mean,
12 SUM06, or AOT40.
13 Table 9-14 presents summaries of ambient O3 exposure patterns in the United States for
14 1982 to 1981 for several indices including the 7-h mean and SUM06. More recent summaries
15 for the entire United States for these indices are not available, but Table 9-15 summarizes more
16 recent data for the central and eastern United States. As shown in Table 9-15, from 1989 to
17 1995, mean 12-h 3-month SUM06 values (in ppm-h) at 41 rural sites in the Clean Air Status and
18 Trends Network were 31.5 for the Midwest, 18.9 for the Upper Midwest, 33.2 for the Northeast,
19 13.2 for the Upper Northeast (NH, ME), 34.5 for the South-Central, and 19.2 for the Southern
20 Peripheral subregions (Baumgardner and Edgerton, 1998). These results are important because
21 these sites were selected to represent rural areas, while many other monitoring sites represent
22 urban or suburban areas. For these same subregions, W126 values ranged from 12.8 to
23 25.6 ppm-h. From 1989 to 1995, O3 concentrations decreased about 5% for daily and 7% for
24 weekly values for most of these sites, after adjusting for meteorological conditions (Holland
25 et al., 1999). These trends were statistically significant at about 50% of the sites (p < 0.05).
26 However, because the trend analysis was intended to examine the efficacy of O3 emissions
27 controls, the trends were adjusted for meteorological conditions. Thus, they do not reflect the
28 actual trends in O3 exposure over time.
29
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3
Species
o
w Bean, cv. Lit
Table 9-20. Ethylene Diurea Effects on Vegetation Responses to Ozone
Description
10-cm pots in OTCs in
Germany
EDU application
Soil drench 200 mL of
150 ppm solution per
plant every 14 days
Ozone exposure Effects of EDU
CF, NF, CF-1 x, CF-2x : O3 reduced pod, shoot, and
mean= 1, 14, 15, 32 ppb root mass. EDU increased
root, leaf, and shoot mass,
but a significant interaction
with O3 occurred only for
root weight.
Reference
Brunschon-Harti
etal. (1995)
Bean, cv. BBL-290
2 expts in 5.5 L pots
in OTCs with 4
O3 treatments
Soil drench every 14
days, in expt 1=0.14,
28, 56, 120 mg/L potting
medium; expt 2 = 0, 8,
16, 32 mg/L
2 expts with CF, NF and
2 constant additions of
O3. 7-h mean O3 (ppb)
for Expt 1 = 34, 70, 95,
121; Expt 2 = 19, 42.
74, 106
Visible injury and reduced
total biomass or yield, even in
CF treatment. Within an O3
treatment, sometimes
increased yield (Expt 2 only).
Miller etal. (1994)
VO
to
Bean
Pots with potting mix at
3 locations in Spain
(2 years at 1 site)
Soil drench of 200 mL of
increasing concentrations
of 100, 150,200,250
ppm every 14 days
(4- 10 mg 1:1 soil)
AOT40 = 0.4-1.8ppm-h
0 to 50% increase in
pod mass, but did not restore
yield at sites with higher O3.
Ribas and Penuelas
(2000)
Bean, cv. Lit
Pots with potting mix at
4 sites in the Netherlands
Soil drench of 200 mL of
increasing concentrations
of 100, 150,200,250
ppm every 14 days
(4- 10 mg 1:1 soil)
AOT40 = 0.64-0.98,
7-h mean = 49-55 ppb
Average 20% yield increase
at all sites.
Tonneijck and
Van Dijk( 1997)
H
6
o
o
H
O
O
HH
H
W
Bean, cv. Lit
Pots with potting mix at
1 site in Belgium
Soil drench of 200 mL of AOT40 = 0.81 ppm-h
increasing concentrations
of 100, 150,200,
250 ppm every 14 days
(4-10 mg 1:1 soil)
16% yield increase.
Vandermeiren et al.
(1995)
Clover, subterranean
Plants in 10-cm pots at
4 rural sites in the
Netherlands for 3 years
100 ml of 150 ppm
solution as soil drench
every 14 days for
2 months
AOT40 = 0-0. 56 ppm-h
for 4-week periods
Injury, but not leaf biomass
was affected by EDU and
O3 exposure.
Tonneijck and
Van Dyk (2002)
-------
to
o
o
Table 9-20 (cont'd). Ethylene Diurea Effects on Vegetation Responses to Ozone
Species
Clover, white
Clover, white, cv. Menna
Description
15-cm pots in field, well
watered, 12 locations
throughout Europe,
3 years
2 expts,10-cmpots in
field in Italy, see also
companion OTC expt
EDU application
100 ml of 150 ppm
solution as soil drench
every 14 days for
3 months
100 ml of 150 ppm
solution as soil drench
every 14 days for
Ozone exposure
AOT40 (28 days) = 0-
20 ppm-h
AOT40 = 15.5,
12.1 ppm«h;
7-h mean = 69, 60
Effects of EDU
Change in biomass ratio, weak
linear relationship (r2 = 0. 16)
stronger relationship using
ANN and climatic factors
n.s.
Reference
Ball etal. (1998)
Fumagalli et al.
(1997)
2 months
vo
to
Clover, white, cv. Menna
Poplar, hybrid
2 expts,10-cm pots in
OTCs in Denmark
Stem injections, field,
cuttings, 1 or 2 years
Soil drench 100 mL of
150 ppm solution every
14 days
Approx. 125 or 250
mg/leaf (low, high
EDU treatments)
5 times every 14 days
CF, NF, NF+ 25,
NF+50 ppb,
O3 exposure not reported
1991:
7-h mean = 56,
AOT40 = 23;
1992:
7-h mean = 59,
AOT40 = 27
No effect of EDU despite
highly significant effect of
O3 on above-ground biomass
No effect on biomass; 6%,
12% more severely O3
damaged leaves in high
EDU for 2 years
(Mortensen and
Bastrup-Birk, 1996)
Ainsworth et al.
(1996)
Pine, loblolly
H
6
o
o
H
O
Radish, cv. Cherry Belle
1 year old half-sib
seedlings in field in TX
for 3 years
Plants in pots in potting
mix exposed for 5 weeks
in southern Sweden.
150, 300, 450 ppm
every 14 days
Soil drench containing
20 mg EDU applied
2 times, 14 days apart
1995, 1996, 1997,
no. h>40ppb= 1723,
2297, 2052;
no. h > 60 ppb = 378,
584, 528;
peak =113, 102, 118
24-h mean = 31 ppb,
7-h mean = 36 ppb,
AOT40 = 1.3 ppm-h.
For EDU 450 trt, above-
ground biomass increased
approx 46% (n.s. in other
treatments
24% increase in hypocotyl
mass, 18% increase in
shoot mass
Manning et al.
(2003)
Pleyeletal. (1999b)
o
HH
H
W
-------
1 9.6.4 Effects of Ozone on Annual and Biennial Species
2 Much of the research on short-lived species during the last decade has been conducted in
3 Europe. Several European studies have focused on wheat with an emphasis on developing
4 critical levels, as discussed below in Section 9.6.4.6.
5 An extensive search of the literature was performed using several electronic databases to
6 identify scientific articles containing quantitative information on both the amount of O3 exposure
7 and its effects on vegetation. Greater emphasis is placed on studies with longer duration with O3
8 exposure concentrations and environmental conditions that were as similar as possible to
9 ambient conditions. Many of the studies reviewed herein were conducted in OTCs. In the
10 United States, nearly all of such studies have used the type of OTC developed by Heagle et al.
11 (1973). For the few studies in the U.S. that used other types of OTCs, they are described briefly
12 in the text. In Europe, a wide variety of styles of OTCs have been used. See Section 9.2 for
13 further information about the use of OTCs. The emphasis in this subsection is on quantifying
14 exposure-response relationships for annual plants, with a focus on the response of above-ground
15 biomass and yield of species grown as crops or occurring as native or naturalized species in the
16 United States. Emphasis is placed on studies not included in the 1996 AQCD (U.S.
17 Environmental Protection Agency, 1996), including a few studies published prior to 1996.
18 However, an attempt is made to compare the results of these recent studies of individual species
19 to those reviewed in the 1996 AQCD.
20
21 9.6.4.1 Effects on Growth, Biomass, and Yield of Individual Species
22 Most research on the effects of O3 on herbaceous species has evaluated growth, biomass, or
23 yield of commercial portions of crop or forage species. It is well established that reproductive
24 organs such as seeds may be particularly sensitive to injury or biomass reductions due to O3, as
25 reviewed recently by (Black et al. 2000}. As discussed in Section 9.4, numerous analyses of
26 experiments conducted in OTCs and with naturally occurring gradients demonstrate that the
27 effects of O3 exposure vary depending on the growth stage of the plant. Plants grown for seed or
28 grain are often most sensitive to exposure during the seed or grain-filling period (Lee et al.,
29 1988; Pleijel et al., 1998; Soja et al., 2000; Younglove et al., 1994), whereas plants grown for
30 biomass production, such as alfalfa, may be sensitive throughout the growth period (Younglove
31 et al., 1994). However, because different species are sensitive during different periods of their
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1 growth and, because planting or germination dates vary throughout large regions even for a
2 single species, no single phenological weighting scheme can appropriately and practically
3 represent all vegetation in all locations in the United States. For natural populations, reductions
4 in seed yield might be particularly important if subsequent seedling establishment is
5 compromised by O3.
6 Green beans (cv. Pros) were grown in pots in OTCs in the Netherlands for 62 days and
7 exposed to 6 treatments consisting of constant O3 additions to charcoal filtered (CF) chambers
8 (see Section 9.2) for 9 h/day (Tonneijck and Van Dijk, 1998). Bean yield response to O3 was
9 nonlinear, with an apparent threshold near the CF30 (charcoal filtered with a constant addition of
10 30 ppb O3) treatment with a 9-h mean O3 concentration of 28 ppb and an AOT40 value of
11 0.1 ppm-h). Yield was reduced by 29% at a 9-h mean value of 44 ppb corresponding to an
12 AOT40 value of 3.6 ppm-h (Table 9-17). Beans were grown in pots in OTCs for 3 months with
13 the following O3 treatments: CF, nonfiltered (NF), CF with O3 added up to ambient, and CF
14 with 2x-ambient O3 (Brunschon-Harti et al., 1995). Ozone reduced pod mass by 56% with a
15 mean concentration of 32 ppb in the 2* ambient treatment as compared with 1 ppb in the CF
16 treatment (daily averaging time not reported). A second treatment factor in this experiment was
17 addition of EDU, as discussed below under the heading "Studies Using Ethylene Diurea as a
18 Protectant". These yield reductions are greater than those previously reported in four similar
19 studies summarized in the 1996 AQCD (Table 5-25 of U.S. Environmental Protection Agency,
20 1996). Greater sensitivity in the more recent experiments may be due to cultivar differences or
21 other differences in experimental protocols.
22 In a study with OTCs on silty loam soil in Beltsville, MD, corn yield was reduced by 13%
23 with exposure to a 7-h mean concentration of 70 ppb O3 compared to a CF treatment with a 7-h
24 mean concentration of 20 ppb (Mulchi et al., 1995; Rudorff et al., 1996a; Rudorff et al., 1996c).
25 In this study, different amounts of O3 were added above ambient levels for 5 days follows: 20,
26 30, 40, 50, 60 ppb, except that O3 was not added to exceed a total concentration of 120 ppb
27 (Rudorff et al., 1996c, 1996a).
28 In two studies conducted in Raleigh, NC, cotton (cv. Deltapine 51) was grown in pots and
29 exposed to CF and 1.5* (nonfiltered, see Section 9.2) O3 in one year, and CF, NF, and
30 1.5 x-ambient O3 in the second year, with ambient and elevated CO2 concentrations (Heagle
31 et al., 1999; Table 9-17). In the first year, yield decreased by 22% with 1.5x-ambient O3 (12-h
January 2005 9-213 DRAFT-DO NOT QUOTE OR CITE
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1 mean value of 71 ppb). In the second year, yield decreased by 21% and 49% with exposure to
2 ambient or 1.5x ambient O3 (12-h mean values of 51 and 78 ppb, Table 9-17). Increased CO2
3 levels prevented or reduced this yield suppression (Heagle et al., 1999). These yield reductions
4 are similar to those reported previously in four similar studies summarized in the 1996 AQCD
5 (Table 5-25 of U.S. Environmental Protection Agency, 1996).
6 In a study of oats in OTCs in southern Sweden, exposure to ambient (NF) O3 did not affect
7 grain yield (Pleijel et al., 1994a). Ambient O3 concentration expressed as a 7-h mean was
8 27 ppb, with only 1 h greater than 80 ppb and none above 90 ppb.
9 The interactive effects of elevated O3 and CO2 additions on potato yield (cv Bintje) were
10 studied in OTCs at 6 sites in northern Europe as part of the CHIP (Changing Climate and
11 Potential Impacts on Potato Yield and Quality) program (Craigon et al., 2002). Ozone was
12 added to a target daily average value of 60 ppb, and AOT40 values across all years and
13 experiments ranged from ~6 to 27 ppm-h. The O3 treatment reduced total tuber yield an average
14 of 4.8% with elevated O3 treatment across all experiments (Craigon et al., 2002). This total
15 effect was statistically significant even though the effects of individual experiments generally
16 were not (Craigon et al., 2002), due to the increased power of the pooled analysis. Several
17 publications report other aspects of the CHIP experiments or present results of individual
18 experiments (De Temmerman et al., 2002a, 2002b; Donnelly et al., 2001a, 2001b; Fangmeier
19 et al., 2002; Finnan et al., 2002; Hacour et al., 2002; Lawson et al., 2002; Pleijel et al., 2002;
20 Vandermeiren et al., 2002; Vorne et al., 2002).
21 The effect of an intermittent constant addition of O3 using a free air exposure system in
22 Northumberland, UK was investigated with the oilseed rape cultivar "Eurol" (Ollerenshaw et al.,
23 1999). Ozone was added for 6 h/day for 17 days. The ambient treatment had a mean value of
24 30 ppb and the O3 addition treatment had a mean of 77 ppb. After overwintering, O3 was added
25 for 32 days for 7 h/day between May and June (mean values of 31 and 80 ppb). Yield was
26 reduced by 14% despite the lack of any foliar symptoms.
27 Field fumigation chambers ventilated with fans on both ends were used to assess effects of
28 five O3 treatments on rice over 3 years in Japan (Kobayashi et al., 1994; 1995). All O3
29 treatments used CF air, and O3 was added to the 0.5, 1.0, 1.5, 2.0, or 2.75* ambient
30 concentration for 7 h/day. Based on a linear regression for the 3 years, yield decreased by 3 to
31 10% at a 7-h mean concentration of 40 ppb (Table 9-17). This decrease is greater than that
January 2005 9-214 DRAFT-DO NOT QUOTE OR CITE
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1 found for rice in earlier studies in California (Kats et al., 1985), although whether this difference
2 is due to differences in cultivars, experimental treatment, or environmental factors cannot be
3 determined.
4 During 3 years in Beltsville, MD, the soybean cultivars Essex and Forrest were exposed to
5 CF air and NF air in OTCs (Chernikova et al., 2000; Robinson and Britz, 2000) with O3 added as
6 described for experiments with corn and wheat at Beltsville (Mulchi et al., 1995; Rudorff et al.,
7 1996c). During 1994 and 1995, as previously found for these cultivars, Essex was less sensitive
8 than Forrest, with yield decreases of 10% (n.s.: p > 0.1) compared to 32% for Forrest (p < 0.01)
9 (Chernikova et al., 2000). There was no evidence of water stress in this experiment. In 1997,
10 the two O3 treatments were CF (7-h mean = 24 ppb) and NF with a constant addition of O3 (7-h
11 mean = 58 ppb) (Robinson and Britz, 2000). The yield of Essex was not significantly affected,
12 while the yield of Forrest was decreased by 21% (Table 9-17).
13 In a study in Raleigh, NC the soybean cultivar Essex was grown in pots and exposed to CF
14 and 1.5* ambient O3 concentrations during three growing seasons (Fiscus et al., 1997). Over the
15 3 years, exposure to an average 12-h mean O3 concentration of 82 ppb reduced soybean yield by
16 41% (Table 9-17). In similar studies also in Raleigh, NC, Essex was exposed to CF, NF, and
17 1.5x ambient O3 for two seasons (Heagle et al., 1998). Yield decreased by 16% and 15% in the
18 2 years by ambient O3 (12-h mean values of 50 and 42 ppb), and decreased by 37 and 40% with
19 exposure to 1.5x ambient O3 (12-h mean values of 79 and 69 ppb, Table 9-17). In this same
20 experiment in the second year, similar yield reductions were observed for the cultivar Holladay,
21 while the growth of cultivar NK-6955 was increased substantially by ambient O3 exposure.
22 All three cultivars were grown in the same chambers in this experiment, and the authors
23 suggested that NK-6955 plants may have shaded the other cultivars to some extent.
24 In a 2-year study using OTCs in Raleigh, NC, the soybean cultivars Coker 6955, Essex,
25 and S53-34' were exposed CF, NF, and 1.5x ambient O3 treatments (Miller et al., 1994).
26 Seasonal mean 12-h O3 concentrations ranged from 14 to 83 ppb. As compared to the CF
27 treatment, ambient O3 exposure (NF treatment) reduced seed yield by 11 to 18% except for
28 Coker 6955 in the first year (1989) which showed a yield increase of 14%. The 1.5x ambient O3
29 treatment reduced yield by 32 to 56% in all cultivars in both years. In a similar subsequent
30 experiment with the cultivar Essex, exposure to a 12-h mean ambient O3 concentration of 50 ppb
January 2005 9-215 DRAFT-DO NOT QUOTE OR CITE
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1 reduced yield by 11%, while exposure to 79 ppb reduced yield by 22% (Table 9-17) (Miller
2 etal., 1998).
3 These yield reductions for soybean are generally similar to those reported previously in
4 13 similar studies summarized in the 1996 AQCD (Table 5-23 of U.S. Environmental Protection
5 Agency, 1996).
6 A reanalysis of 7 years of data from OTC experiments with wheat in Ostad, Sweden
7 showed that relative yield linearly decreased with increasing O3, with a maximum yield loss of
8 23% at an AOT40 value of 15 ppm-h (Danielsson et al., 2003). A very similar response was
9 found using the flux (stomatal conductance) model of Emberson et al. (2000b), and a similar
10 amount of the variance was explained by the flux model (for AOT40 model, r2 = 0.34 and for the
11 Emberson flux model, r2 = 0.39). A modified flux model developed and calibrated for this site
12 also had a similar linear response equation, but explained much more of the variance (r2 = 0.90).
13 During the 1990s, a major European research program investigated the combined effects of
14 CO2, O3, and other physiological stresses on wheat (Bender et al., 1999; Hertstein et al., 1996,
15 1999; Jager et al., 1999). The European Stress Physiology and Climate Experiment ("ESPACE-
16 wheat") program included 13 experiments in OTCs at eight sites in northern Europe over 3
17 years. Low-and high-O3 exposures in these experiments had the following values: 12-hmean
18 (SD) low = 26.3 ppb (12.2), 12-h mean (SD) high = 51.37 (18.3) ppb, AOT40 mean (SD) low =
19 6.2 (8,5) ppm-h, AOT40 mean (SD) high = 28.3 (23.0) ppm-h, as calculated from data presented
20 in Table 3 of Hertstein et al. (1999). An analysis of all 13 experiments showed that high O3 at
21 ambient CO2 reduced yield by 13% on average (Bender et al., 1999). However, this reduction
22 was not statistically significant based on an ANOVA, and the authors concluded that the wheat
23 cultivar Minaret may be relatively tolerant to O3 (Bender et al., 1999). Results of some
24 individual studies within this program have been reported previously (Donnelly et al., 1999;
25 Fangmeier et al., 1996, 1999; Mulholland et al., 1997, 1998b, 1998a; Pleijel et al., 2000b).
26 In a study with OTCs on silty loam soil in Beltsville, MD, wheat yield was reduced by
27 20% on average over 2 years with 7-h mean concentrations of 61 and 65 ppb O3 compared with
28 CF treatment with a 7-h mean concentration of 20 ppb (Mulchi et al., 1995; 1996a; Rudorff
29 et al., 1996c). In the above study, different amounts of O3 were added above ambient (levels for
30 5 days) as follows: 20, 30, 40, 50, 60 ppb, except that O3 was not added to exceed a total
31 concentration of 120 ppb (Rudorff et al., 1996c). Wheat grown in pots in OTCs was exposed to
January 2005 9-216 DRAFT-DO NOT QUOTE OR CITE
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1 elevated O3 and water stress in Germany, and yield was decreased by 35% in the 2x -ambient
2 treatment with a 7-h mean O3 concentration of 71 ppb statistically significant effects were not
3 seen in the 1 x-ambient treatment (Fangmeier et al., 1994). In two studies conducted in Raleigh,
4 NC, soft red winter wheat was grown in pots and exposed to CF, NF, and 1.5x-ambient O3, with
5 ambient and elevated CO2 concentrations (Table 9-17) (Heagle et al., 2000). In the first
6 experiment, eight cultivars were exposed to 12-h mean O3 concentrations of 27, 47, and 90 ppb,
7 and in the second experiment two of these cultivars were exposed to 22, 38, and 74 ppb. There
8 was a trend toward decreased yield in both experiments, but these trends were not statistically
9 significant. The wheat cultivar Drabant was exposed to CF, NF, and a constant addition of
10 35 ppb during 1992 and 1993 using the Heagle-type OTCs (Heagle et al., 1973) in Finland
11 (Ojanpera et al., 1998). The following 12-h mean O3 exposures were observed in 1992: 14, 30,
12 61 ppb. In 1993, the values were 9, 21, and 45 ppb, (see Table 9-17 for AOT40 values and other
13 information). Yield was reduced 13% in each year by the added O3 treatment.
14 The effect of an intermittent constant addition of O3 using a free air exposure system was
15 investigated with the winter wheat cultivar Riband in Northumberland, UK (Ollerenshaw and
16 Lyons, 1999). Ozone exposures expressed as AOT40 values for September and October 1992
17 were 0.14 and 3.5 ppm-h; while for April to August 1993, values were 3.5 and 6.2 ppm-h. Yield
18 was reduced by 13%.
19 These results provide an additional line of evidence supporting the OTC-studies that
20 demonstrated yield reductions in wheat due to O3 exposures that occur in the United States.
21 These yield reductions for wheat are generally similar to those reported previously in
22 22 comparable studies summarized in the 1996 AQCD (Table 5-25 of U.S. Environmental
23 Protection Agency, 1996).
24
25 9.6.4.2 Effects on Plant Quality
26 In addition to reductions in biomass or crop yield, O3 may also reduce the quality or
27 nutritive value of annual species. Many studies have shown effects of O3 on various measures of
28 plant organs that affect quality, with most studies focusing on characteristics important for food
29 or fodder (U.S. Environmental Protection Agency, 1996).
30 The effect of a continuous intermittent addition of O3 using a free air exposure system in
31 Northumberland, UK was investigated with the oilseed rape cultivar Eurol as discussed above
January 2005 9-217 DRAFT-DO NOT QUOTE OR CITE
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1 (Ollerenshaw et al., 1999). Ozone exposures expressed as AOT40 values for August to October
2 1991 were 0.2 and 3.8 ppm-h and for June 1992 were 0.7 and 8.1 ppm-h. Yield quality measured
3 as crude protein and oil content was decreased significantly. Because the price of the product is
4 reduced in direct proportion to the oil content, such a decrease represents a substantial loss to
5 growers (Ollerenshaw et al., 1999).
6 Two wheat cultivars, Massey and Saluda, were each grown for one year each in Beltsville,
7 MD (Table 9-17) and exposed to either CF or an addition of 40 ppb for 7 h/day for 5 days/week
8 (Mulchi et al., 1995, 1996a; Rudorff et al., 1996c). Milling and baking quality scores and flour
9 protein were not significantly affected by elevated O3 exposure, but the softness equivalent was
10 increased slightly (2.4%) in both experiments (Rudorff et al., 1996b). The authors concluded
11 that these changes, along with other slight changes due to an increased CO2 treatment, suggested
12 that O3 and CO2 had only minor effects on wheat grain quality. In wheat grown in Sweden, the
13 harvest index was significantly decreased and the protein content increased due to exposure to a
14 12-h mean of 48 ppb (Gelang et al., 2000). In an analysis of 16 experiments conducted with
15 spring wheat and either O3 or CO2 exposures in four Nordic countries, a negative linear
16 relationship was found between grain yield and grain protein content (y = -0.38* +138.6,
17 expressed as percentages of the NF treatment (Pleijel et al., 1999a).
18 For three soybean cultivars grown in Raleigh, NC, O3 significantly decreased oleic acid
19 content, although the authors stated that the reduction was not large enough to be economically
20 important (Heagle et al., 1998).
21 In a UK study, potato exposed during 1998 to an AOT40 value of 12.5 ppm-h in OTCs
22 (in Nottingham) resulted in the paste from tubers being more viscous (Donnelly et al., 2001b).
23 In this study, an AOT40 exposure of 27.11 ppm-h in 1999 caused starch granules to be less
24 resistant to swelling, and total glycoalkaloid content was increased due to an increase in a-
25 solanine (Donnelly et al., 200Ib). Such increases in glycoalkaloid content have been observed
26 previously in potato (Pell and Pearson, 1984) and may be important, because glycoalkaloids
27 cause bitter flavors and, at higher concentrations, cause toxicity. The authors indicated that
28 levels found in this study approached those that may cause bitterness, but not those of concern
29 for toxicity (Donnelly et al., 2001b).
30 In the CHIP program the effects of O3 were studied using OTCs at six sites in northern
31 Europe, and yield decreases were observed as described above. The reducing sugar and starch
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1 content of tubers decreased linearly due to O3 exposure, while the ascorbic acid concentration
2 increased linearly (Vorne et al., 2002). Compared to the CF treatment, exposure to an AOT40
3 value of 14 ppm-h decreased starch concentrations by 2%, decreased reducing sugar
4 concentration by 30%, and increased ascorbic acid concentration by 20%. While the changes in
5 reducing sugars and ascorbic acid increase tuber quality, the reduction in starch concentration
6 decreases tuber quality.
7 In two 1-year studies using OTCs in commercial fields in Spain, the soluble solids content
8 of watermelon was decreased 4 to 8% due to seasonal O3 exposures as follows: AOT40 = 5.96
9 ppm-h and SUM06 = 0.295 ppm-h in one year; and AOT40 = 18.9, SUM06 = 4.95 in the second
10 year (Gimeno et al., 1999).
11
12 9.6.4.3 Effects on Foliar Symptoms
13 For most annual crop species, the most important effects of O3 are on yield of the
14 commercially important part of the crop, expressed as the mass of the harvested portion.
15 However, for some crops, foliar symptoms are important if they reduce the marketability of the
16 crop. This is why efforts have been made to identify O3 exposures associated with foliar
17 symptoms. In Europe, Level I critical levels have been determined for such effects based on
18 observations from experiments conducted in 15 countries under the auspices of the United
19 Nations Economic Commission for Europe International Cooperative Programme on effects of
20 air pollution and other stresses on crops and non-woody plants (UN/ECE ICP-Vegetation;
21 formerly ICP-Crops), as well as on observations of symptoms in commercial fields from 1993 to
22 1996 (Benton et al., 1995, 2000). Because the occurrence of symptoms increased with greater
23 humidity, these levels took into account the vapor pressure deficit (VPD). Two short-term
24 critical levels were derived from 1995 data: an AOT40 value of 0.2 ppm-h over 5 days when
25 mean VPD is below 1.5 KPa (0930 - 1630 h), and a value of 0.5 ppm-h when the mean VPD is
26 above 1.5 Kpa (Benton et al., 1996). The 1996 data supported the critical levels in 83% of
27 observations, although symptoms occurred on three occasions when the AOT40 was less than
28 0.05 ppm-h and the VPD was very low — less than 0.6 Kpa. The authors concluded that these
29 critical levels are a good indicator of the likelihood of foliar symptoms, but that further
30 refinement may be required, such as including factors that modify O3 uptake by stomata.
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1 In a study in Germany, 25 native herbaceous species were exposed to several square-wave
2 O3 exposures in CF OTCs (Bergmann et al., 1999). Six of the 25 species showed O3-specific
3 symptoms, and five species responded to single-day peaks of a single day. The most sensitive
4 species exhibiting O3-specific symptoms were Cirsium arvense and Sonchus asper, which
5 responded to AOT40 values < 1.5 ppm-h (Bergmann et al., 1999).
6 In the United States, attention has been paid to foliar symptoms on annual plants in
7 previous criteria documents, and most recent literature has focused on other topics. However, a
8 study of O3 effects was undertaken from 1990 to 1993 in Acadia National Park in Maine, which
9 experiences elevated O3 exposures due to transport from urban areas located upwind (Kohut
10 et al., 2000). Because this study examined both herbaceous perennial species and woody
11 perennial species, the results are discussed in the perennial section below. Other studies and
12 biomonitoring programs in other U.S. National Parks and over larger areas are oriented primarily
13 towards forested ecosystems, so such studies are also discussed in the perennial subsection
14 below as well as in Section 9.7.
15
16 9.6.4.4 Other Effects
17 In addition to the effects seen on individual species grown in monocultures, effects may
18 occur due to competition when species are grown in mixtures, as reviewed by Davison and
19 Barnes (1998). Such effects have been investigated primarily in species grown for forage, which
20 often include perennial species; therefore, this topic is discussed in the herbaceous perennial
21 section, Section 9.6.5.1. Other stresses may interact with O3, including pathogens and pests as
22 well as environmental conditions such as drought. These issues are discussed in Section 9.4.
23 Effects on reproduction are discussed briefly below.
24 Several studies during recent decades have demonstrated O3 effects on different stages of
25 reproduction. Effects of O3 have been observed on pollen germination, pollen tube growth,
26 fertilization, and abortion of reproductive structures as reviewed by Black et al. (2000). This
27 issue is not addressed here because reproductive effects will culminate for seed-bearing plants in
28 seed production, and the substantial body of evidence relating O3 exposure and reduced seed
29 production is discussed above. However, one example of a native species will be presented
30 because of its implications for extrapolating exposure-response data to noncommercial species.
31 Spreading dogbane has been identified as a useful species for O3 biomonitoring because of
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1 O3-induced diagnostic symptoms (Kohut et al., 2000). A study in Massachusetts found that
2 exposure to O3 in NF OTCs or ambient plots for 103 days produced significantly fewer flowers,
3 and also that fewer of these flowers survived to produce mature fruits (Bergweiler and Manning,
4 1999). Because foliar symptoms were not common, the authors concluded they are not required
5 for effects on reproduction to occur. Genotoxic effects and effects on population genetics are
6 discussed in Section 9.4.
7
8 9.6.4.5 Scaling Experimental Data to Field Conditions
9 Substantial effort has been invested in the design of OTCs for assessing the effects of air
10 pollutants on vegetation under near-ambient conditions. The design, construction, and
11 performance of many types of chambers has been reviewed extensively (Hogsett et al., 1987a,
12 1987b). Despite such design efforts, the influence of experimental chambers on exposure-
13 response functions has been debated for many years (e.g., Manning and Krupa, 1992) because
14 several factors differ between OTC studies and actual fields. This issue is addressed in Section
15 9.2, and only a few comments about the implications of chamber artifacts for interpreting
16 exposure-response relationships are presented here.
17 While it is clear that chambers can alter some aspects of plant growth, the more important
18 issue is whether they alter the response of plants to O3. A review of such chamber studies done
19 in California found that plants responded similarly to O3 whether OTCs, closed-top chambers, or
20 air exclusion systems were used; differences were found for fewer than 10% of growth
21 parameters (Olszyk et al., 1986). In a different review of literature about Heagle-type OTCs
22 (Heagle et al., 1988), the authors concluded that "Although chamber effects on yield are
23 common, there are no results showing that this will result in a changed yield response to O3."
24 A more recent study of chamber effects examined the responses of tolerant and sensitive white
25 clover clones to ambient O3 in greenhouse, open-top, and ambient plots (Heagle et al., 1996).
26 For individual harvests, O3 reduced the forage weight of the sensitive clone 7 to 23% more in the
27 greenhouse than in OTCs. However, the response in OTCs was the same as in ambient plots.
28 Several studies have shown very similar yield response to O3 for plants grown in pots or in the
29 ground, suggesting that even such a significant change in environment does not alter the
30 proportional response to O3, at least as long as the plants are well watered (Heagle, 1979; Heagle
31 etal., 1983).
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1 Most experiments investigating the O3 effects on annual vegetation provide adequate water
2 to avoid substantive drought stress. Because drought stress has generally been shown to reduce
3 the effect of O3 on annual vegetation, such experiments may tend to overestimate O3 effects on
4 crops and especially on unmanaged or seminatural vegetation.
5 As mentioned above, the use of O3 flux rather than exposure is theoretically more realistic,
6 and such an approach would also address the vertical gradient issue (Section 9.2). Recently, a
7 number of investigators have suggested that modeling O3 flux can improve estimates of O3
8 effects on vegetation. Models of O3 flux can reduce the variation in the response to O3 that is
9 sometimes observed between years in an experiment (Emberson et al., 2000a, 2000b; Fuhrer
10 et al., 1992; Griinhage et al., 1993; Griinhage and Haenel, 1997; Pleijel et al., 2000a). In a study
11 of O3 deposition to an oat crop in OTCs, O3 flux in the chamber was estimated to be up to twice
12 that in an adjacent field based on a K-theory approach and measurements of stomatal
13 conductance and environmental conditions (Pleijel et al., 1994b). These measurements were
14 made for 2 hours on 5 days when the canopy was physiologically active and wind speeds were
15 moderate. However, the O3 flux in a chamber without plants was nearly as high as that in the
16 open field. The authors conclude that O3 uptake in the chamber was between 100 and 200% of
17 that in the field. These models of flux have a sound biological and meteorological basis and are
18 useful for interpreting experimental data. Flux models have been successfully applied at
19 intensive study sites with detailed site-specific data on stomatal conductance and
20 micrometeorological conditions (e.g., Griinhage et al., 1993b, 1994; Fredericksen et al., 1996).
21 Yet even at a single well-studied site, different methods can provide different estimates of O3
22 flux. For example, at a site in a vineyard in California, an evapotranspiration-based method
23 overestimated the O3 flux as compared to an eddy covariance approach by 20 to 26% (Massman
24 and Grantz, 1995). At a site in a nearby cotton field the evapotranspiration-based approach
25 overestimated the eddy-covariance method by 8 to 38%.
26 Interest has been increasing in recent years, particularly in Europe, in using mathematically
27 tractable flux models for O3 assessments at the regional and national scale (Emberson et al.,
28 2000a, 2000b). However, methods for scaling site-specific models of O3 flux to large areas
29 remains an active and challenging area of research (Massman et al., 1994; Massman and Ham,
30 1994; Wesely and Hicks, 2000). Reducing uncertainties in flux estimates for areas with diverse
31 surface or terrain conditions to within ±50% requires "very careful application of dry deposition
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1 models, some model development, and support by experimental observations" (Wesely and
2 Hicks, 2000). As an example, the annual average deposition velocity of O3 among three nearby
3 sites in similar vegetation was found to vary by ±10%, presumably due to terrain (Brook et al.,
4 1997). Moreover, the authors stated that the actual variation was even greater, because stomatal
5 uptake was unrealistically assumed to be the same among all sites, and flux is strongly
6 influenced by stomatal conductance (Brook et al., 1997). Stomatal conductance is affected by
7 factors such as temperature, vapor pressure deficit, plant water status, and temperature. It is a
8 challenging task to obtain such data for regional, national, or continental flux modeling (Pleijel,
9 1998). This topic is addressed further in Section 9.5.
10 If a flux approach is not used, another issue in scaling from experimental data to national
11 assessments is the selection of an O3-exposure index. As discussed in Section 9.5, there is no
12 evidence to support a single value for defining a cutoff value for calculating a peak-weighted O3
13 index. In Europe, it has become common practice to use a cutoff value of 40 ppb, as in the
14 AOT40 index; while in the United States, a cutoff value of 60 ppb has been used, as in the
15 SUM06 index. The W126 index has been promoted, because it has a continuous weighting
16 function without an arbitrary cutoff value; but few exposure-response studies have used this
17 index to date. There is evidence from some European studies that a cutoff value either lower
18 (Pleijel et al., 1997) or higher (Finnan et al., 1996, 1997) may provide a better statistical fit to
19 experimental data. However, the choice of the type of weighting depends on the type of
20 response model that is fit. In an evaluation of numerous indices fit to seven studies of the
21 response of spring wheat to O3 in OTCs in Northern Europe, different types of indices performed
22 better for linear response models than for Wiebull response models (Finnan et al., 1997). In this
23 study, cumulative indices that give greater weight to higher O3 concentrations performed best;
24 and, while the AOT40 was not the best index, it did perform well, and slightly better than the
25 SUM06 in conjunction with a linear model (Finnan et al., 1997). Although this section is
26 focused on annual plants, the issue of cutoff values applies to perennial species as well;
27 therefore, two examples of tree species will be discussed here. Aspen seedlings were exposed to
28 O3 in controlled environment chambers for 4 weeks, and a significantly greater degree of foliar
29 symptoms occurred with a pattern of variable peak exposures compared to a constant peak
30 exposure with the same total SUM06 exposure (Yun and Laurence, 1999b). However, a field
31 experiment in large OTCs in Ithaca, NY with these two clones exposed to ambient, 1.7*,- and
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1 3 x-ambient O3 concentrations (SUM06 = 1, 20, and 62 ppb) found SUM06 to be highly
2 correlated with foliar symptoms and biomass effects (Yun and Laurence, 1999a). When silver
3 birch was exposed to different patterns of O3 with the same AOT40 value in growth chambers,
4 high peak concentrations and exposure shape were found to be important for symptom
5 production (Oksanen and Holopainen, 2001). However, growth reductions were best predicted
6 by total cumulative exposure.
7 An investigation of different exposure indices was conducted using data from experiments
8 conducted during 1993 at eight sites in the eastern United States in which O3 sensitive and
9 O3-tolerant white clover genotypes were grown using methods developed by Heagle et al.
10 (1994). A statistical approach based on profile likelihoods was used to estimate parameters in
11 generalized exposure indices similar to the SUM06 and AOT40 indices (Blankenship and
12 Stefanski, 2001). The results showed that for the SUMX family of indices, where X is a cutoff
13 value, hourly O3 concentrations over -71 ppb contribute the most to yield prediction. For the
14 AOTX family of indices, the parameter was 54.4. These values are similar to those used in the
15 SUM06 and AOT40 indices already in use. Furthermore, investigation of weighting for time of
16 day confirmed the importance of the mid-afternoon hours for this data set, unlike the results
17 found for wheat in Sweden (Danielsson et al., 2003; Pleijel et al., 2000a).
18
19 9.6.4.6 European Critical Levels
20 During the late 1980s and 1990s, substantial effort was expended in Europe to quantify the
21 effects of O3 on crops and trees. A focus for this research effort was to develop and refine
22 "critical levels" for O3 under the auspices of the United Nations Economic Commission for
23 Europe (UNECE) Convention on Long-Range Transboundary Air Pollution (Sanders et al.,
24 1995). A critical level was defined as "the concentration of pollutant in the atmosphere above
25 which direct adverse effects on receptors , such as plants, ecosystems or materials, may occur
26 according to present knowledge" (United Nations Economic Commission for Europe, 1988).
27 Areas where the levels were exceeded were identified to plan abatement strategies. Critical
28 levels are not air quality standards as are used in the United States for criteria pollutants, nor are
29 they legally enforced. However, they have been used successfully for planning reductions in
30 sulfur and nitrogen pollution. Critical levels for O3 are intended to prevent long-term deleterious
31 effects on the most sensitive plant species under the most sensitive environmental conditions, but
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1 not to quantify O3 effects. The nature of the "significant harmful effects" is not specified in the
2 original definition, which allows different levels to be defined for different effects, such as foliar
3 symptoms or reduction in crop yield. Different levels also have been set for crops, forests and
4 seminatural vegetation. The caveat, "according to present knowledge" is important, because
5 critical levels are revised periodically as new scientific information becomes available.
6 Because these critical levels were designed for use in regional modeling and mapping, they
7 were required to be of a simple form (Sanders et al., 1995). During the early 1990s, the AOT40
8 index was selected as the form of the critical level, and Level I critical levels were determined
9 for arable crops and forestry. The AOT40 index is defined as the sum of the difference between
10 the hourly concentration (in ppb) and 40 ppb when the concentration exceeds 40 ppb for the
11 hours when global radiation exceeds 50 W nT2 (Fuhrer et al., 1997). In regression analysis of
12 15 OTC studies of spring wheat, including one study from the United States and 14 from
13 locations ranging from southern Sweden to Switzerland, an AOT40 value of 5.7 ppm-h was
14 found to correspond to a 10% yield loss, and a value of 2.8 ppm-h corresponded to a 5% yield
15 loss (Fuhrer et al., 1997). Because a 4 to 5% yield loss could be detected with a confidence level
16 of 99%, a critical level of 3 ppm-h was selected an AOT40 value of in 1996 (Karenlampi and
17 Skarby, 1996). This critical level is defined for a 3-month period calculated for daylight hours.
18 It was suggested that a 5-year average should be used to assess the exceedance of this critical
19 level, because this value was associated with a small yield decrement (Fuhrer et al., 1997). This
20 value is currently used for all crops, because it is the best-supported value and because the
21 limited data from other crop species do not provide strong evidence that a more stringent value is
22 required (Fuhrer et al., 1997).
23 The Level I critical levels were successfully used in the 1990s to map areas of exceedance,
24 but the research led to the conclusion that simple, exposure-based levels lead to overestimation
25 of the effects in some regions and underestimation in others. The main problem is that other
26 environmental factors (e.g., vapor pressure deficit, water stress, temperature, light) alter O3
27 uptake and its effects. Therefore, the decision was made to work towards a flux-based approach,
28 aiming to be able to model O3 flux-effect relationships for the three vegetation types (i.e., crops,
29 forests, seminatural vegetation). Progress has been made in modelling flux under experimental
30 conditions (Danielsson et al., 2003; Griinhage et al., 2001; Griinhage and Jager, 2003; Pleijel
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1 et al., 2000a), but because the UNECE meeting in November 2002 concluded that more research
2 is needed, the subject is to be reassessed in 2 to 5 years' time.
3 There has been criticism that the Level I critical level for crops overestimates O3 effects in
4 the Mediterranean countries, because it was developed based on studies in Northern Europe
5 (De Santis, 1999, 2000). However, there is evidence of substantial crop loss due to O3 in some
6 southern European counties, such as the Po valley in Northern Italy (Fumagalli et al., 2001).
7 In these studies Heagle type OTCs were used. Losses in NF chambers as compared to CF
8 chambers over several years at two sites ranged from 11.2 to 22.8% for barley and wheat, from
9 0.3 to 31.5% for other crop species, and from 4.1 to 19.8% for forage species (Fumagalli et al.,
10 2001). Surprisingly, the least effect was observed for soybean, despite AOT40 values of 9.32
11 ppm-h, 3 x the Level I critical level. Similarly, a review of studies in Northern Italy found that
12 ambient O3 episodes have been reported to cause foliar symptoms on 24 agricultural and
13 horticultural crops in commercial fields (Fumagalli et al., 2001). Ambient O3 has also been
14 reported to cause yield losses in several crop species, although no data on O3 exposure were
15 presented by the reviewers (Fumagalli et al., 2001).
16 The Level I approach has also been criticized for focusing only on a single annual species
17 (wheat) and a single woody perennial species (beech). However, this species focus is
18 appropriate, because the goal was to determine an exposure-response relationship for a sensitive
19 species based on available data. In support of standards in Germany, an effort was made to
20 combine data from different species, and consisted of a meta-analysis of studies conducted in
21 both closed chambers and OTCs (Griinhage et al., 2001). In this study, experiments published
22 between 1989 and 1999 were analyzed if they met the following conditions: (1) a significant O3
23 effect was determined (2) exposure conditions were well defined, (3) foliar symptoms, growth,
24 or yield was measured, and (4) plant species were relevant to Europe (Griinhage et al., 2001).
25 Despite the focus on European species, many of the species studies also occur in the United
26 States. Separate regressions for herbaceous plants and for tree species were created as a function
27 of duration of exposure at a given level of O3 exposure at the top of the plant canopy. These
28 regression equations, with confidence limits and with correction for the vertical gradient in O3
29 from the top of the quasi-laminar boundary layer, can be used to define whether effects are
30 unlikely (below the lower confidence limit), probable (between the confidence limits), or highly
31 likely (above the upper confidence limit) to occur near a given O3-monitoring station.
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1 A further concern about the Level 1 approach is that foliar symptoms, rather than biomass,
2 may be an important endpoint, because foliar symptoms may be more sensitive than is biomass
3 increment (VanderHeyden et al., 2001). In an OTC study in southern Switzerland, it was shown
4 that a number of tree species show foliar symptoms at AOT40 values lower than the Level 1
5 value of 10 ppm-h (VanderHeyden et al., 2001).
6
7 9.6.4.7 Summary of Effects on Short-Lived Species
8 For annual vegetation, the data summarized in Table 9-17 show a range of growth and
9 yield responses both within species and among species. Nearly all of these data were derived
10 from studies in OTCs, with only two studies using open-air systems in the UK (Ollerenshaw
11 et al., 1999; Ollerenshaw and Lyons, 1999). It is difficult to compare studies that report O3
12 exposure in different indices such as AOT40, SUM06, or 7-h or 12-h mean values. However,
13 when such comparisons can be made, the results of this recent body of research confirm earlier
14 results summarized in the 1995 AQCD (U.S. Environmental Protection Agency, 1996).
15 A summary of earlier literature concluded that a 7-h, 3-month mean of 49 ppb corresponding to a
16 SUM06 exposure of 24.4 ppm-h would cause 10% loss in 50% of 49 experimental cases (Tingey
17 et al., 1991). A similar study using a 24-h, rather than 7-h, averaging period found that a
18 SUM06 exposure of 26.4 ppb would cause 10% loss in 50% of 54 experimental cases (Lee et al.,
19 1994a, 1994b). Recent data summarized in Table 9-17 support this conclusion. These values
20 represent ambient exposure patterns that occur in some years over large portions of the United
21 States. Some annual species such as soybean are more sensitive, and greater losses may be
22 expected (Table 9-17). Thus, the recent scientific literature supports the conclusions of the 1996
23 AQCD that ambient O3 concentrations are reducing the yield of major crops in the United States.
24 Much research in Europe has used the AOT40 exposure statistic, and substantial effort has
25 gone into developing Level-1 values for vegetation. Based on regression analysis of 15 OTC
26 studies of spring wheat, including one study from the United States and 14 from locations
27 ranging from southern Sweden to Switzerland, an AOT40 value of 5.7 ppm-h was found to
28 correspond to a 10% yield loss, and a value of 2.8 ppm-h corresponded to a 5% yield loss
29 (Fuhrer et al., 1997). Because a 4 to 5% decrease could be detected with a confidence level of
30 99%, a critical level of an AOT40 value of 3 ppm-h was selected in 1996 (Karenlampi and
31 Skarby, 1996).
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1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
In addition to reductions in crop yield, O3 may also reduce the quality or nutritive value of
annual species. Many studies have shown effects of O3 on various measures of plant organs that
affect quality, with most studies focusing on characteristics important for food or fodder. These
studies indicate that there may be economically important effects of ambient O3 on the quality of
crop and forage species. Previous criteria documents have concluded that foliar symptoms on
marketable portions of crops and ornamental plants can occur with seasonal 7-h mean O3
exposures of 40 to 100 ppb (U.S. Environmental Protection Agency, 1978, 1986, 1996). The
recent scientific literature does not refute this conclusion.
The use of OTCs may reverse the usual vertical gradient in O3 that occurs within a few
meters above the ground surface (Section 9.2). Such a reversal suggests that OTC studies may
overestimate, to some degree, the effects of an O3 concentration measured several meters above
the ground. However such considerations do not invalidate the conclusion of the 1996 AQCD
(U.S. Environmental Protection Agency, 1996) that ambient (Table 9-14, Table 9-21) O3
concentrations are sufficient to reduce the yield of major crops in the United States.
Table 9-21. Ozone Exposures at 35 Rural Sites in the Clean Air Status and Trends
Network in the Central and Eastern United States From 1989 to 1995
Subregion
Midwest
Upper
Midwest
Northeast
Upper
Northeast
South
Central
Southern
Periphery
SUM06
12-h, 3-Month
Mean
31.5
18.9
33.2
13.2
34.5
19.2
UM06
12-h,
3-Month SD
10.2
8.5
11.9
8.6
16.6
7.6
W126
3-Month
Mean
25.1
16.0
26.6
12.8
25.6
15.2
W126
3-Month
SD
7.7
5.9
9.5
6.5
11.5
5.4
Max.
8h > 80 ppb(n)
Mean
13.8
5.6
15.8
3.3
7.1
1.9
Max.
8h > 80 ppb(n)
SD
10.6
5.6
12.2
5.5
10.0
1.6
Units for SUM06 and W126 are ppm-h.
Source: Baumgardner and Edgerton (1998).
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1 9.6.5 Effects of Ozone on Long-Lived (Perennial) Species
2 Although there has been considerable research in Europe on annual species during the past
3 10 years, much recent research in the United States has focused on perennial species. In Europe,
4 and in a few studies in the United States, effects of O3 on mixtures of annual and perennial
5 herbaceous species have been investigated using growth chambers, greenhouses, and OTCs.
6 Section 9.6.5.1 reviews such studies, with an emphasis on studies using OTCs.
7
8 9.6.5.1 Herbaceous Perennial Species
9 Two alfalfa cultivars were grown in pots and exposed to CF, NF, 1.5 x -ambient and
10 2x-ambient O3 concentrations in two 1-year studies in Quebec, Canada (Renaud et al., 1997).
11 One cultivar, Apica, is commonly grown in the region, and another, Team, is normally grown
12 farther south and is more tolerant to O3. For Apica in both years and for Team in 1991, O3
13 exposure caused a linear reduction in above-ground biomass. In the NF treatment, growth of
14 Apica was decreased by 31 and 21% in the 2 years, while the growth of Team was reduced by
15 14% in 1991, but not reduced in 1992. The authors suggested that the differing effects on Team
16 could be due to different progenies and propagation methods in the 2 years or to more rapid
17 growth in 1991 along with higher O3 peak values in 1991. In 1991, O3 maxima exceeded 60 ppb
18 in 15 days, whereas in 1992 there were only three such days. At the end of the growing season,
19 total starch reserves in roots were decreased by O3 due primarily to a decrease in root mass,
20 which the authors suggested could accelerate decline in alfalfa yields. These yield reductions are
21 generally similar to those reported previously in five similar studies summarized in the 1996
22 AQCD (Table 9-25 of U.S. Environmental Protection Agency, 1996).
23 A study in Alabama exposed early- and late-season-planted bahia grass (cultivar
24 Pensacola) in OTCs to CF, NF or 2x-ambient O3 treatments (Muntifering et al., 2000). Ozone
25 exposures expressed as 12-h mean values over the 24-week experiment were 22, 45, and 91 ppb,
26 and the highest ambient O3 concentrations were recorded in late June, late July, late August and
27 mid-September at approximately 90 ppb. Above-ground biomass growth was reduced by the NF
28 treatment for the first and second harvest by 34% and 29% for the early-season planting, but
29 statistically significant effects were not observed in the late-season planting (Table 9-18). The
30 2x-ambient treatment did not cause further significant reductions in biomass. The authors
31 suggested that the lack of a significant O3 effect in the late planting may have been due to the
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1 shorter total O3 exposure time as well as to the lower O3-exposure concentrations during the
2 weeks immediately preceding harvest. These results are important, because this is an
3 economically important species and because previous studies have focused on grass species that
4 use the C3, rather than C4, metabolic pathway.
5 An investigation of the use of different O3 indices and averaging times on the correlation
6 with growth effects was undertaken with the NC clover system (Heagle et al., 1995). For 2 years
7 of data at six sites in Massachusetts, Oregon, North Carolina, California (2 sites), and Virginia,
8 averaging time was found to be more important than the choice of the type of index including
9 mean, SUM06, and AOT40 (Heagle and Stefanski, 2000). The best correlation between O3
10 exposure and the ratio of sensitive to tolerant clover types was found for the 6-h period from
11 1000 to 1600 h. For this period, very similar r2 values (0.91 to 0.94) were found for SUM06,
12 W126, and AOT40 (Heagle and Stefanski, 2000). For the above indices, a linear relationship
13 was found, with no effect in Corvallis, OR with exposure to a SUM06 value of 10.2 ppm-h and a
14 ratio of 0.53 (sensitive/tolerant) at San Bernardino with a SUM06 exposure of 39.4 ppm-h
15 (Heagle and Stefanski, 2000).
16 In a study of the biomass ratio of O3-sensitive versus O3-insensitive clover at 14 sites in
17 Europe during 1996 to 1998, a model that was developed using artificial neural network (ANN,
18 see Section 9.2) techniques had r2 values for the training data of 0.84 and for unseen validation
19 data of 0.71 (Mills et al., 2000). The predictive factors in the model were AOT40, 24-h mean
20 O3, daylight mean temperature, and 24-h mean temperature. This model was selected after a
21 thorough investigation of a number of models using many more or fewer parameters using both
22 ANN and multiple linear regression techniques. This model predicted that a 5% reduction in
23 biomass ratio was associated with AOT40 values in the range 0.9 to 1.7 ppm-h accumulated over
24 28 days, with plants being most sensitive under conditions of low NOX, moderate temperature,
25 and high 24-h mean O3 concentration.
26 Two experiments in Ontario investigated effects of square-wave additions of O3 for 7 h on
27 1 day/week for 7 weeks on the growth and fruit production of two cultivars of primocane (first-
28 year fruiting) raspberry plants (Sullivan et al., 1994). Plants were grown in OTCs and exposed
29 to ambient O3 (mean values of < 4 ppb). In the both experiments, additions of 12 ppb O3 on
30 1 day/week for 7 weeks did not significantly reduce fruit yield. In the second experiment,
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1 additions of 24 ppb significantly reduced the fruit yield 52% below that of the ambient treatment
2 in one cultivar (Heritage), but not in another (Redwing).
3 Exposure to a square-wave 8-h mean O3 concentration of 92 ppb for 62 days in an
4 experiment in OTCs the UK did not significantly reduce the total yield of strawberry fruits, but
5 did decrease the average size of the fruits by 14% (Drogoudi and Ashmore, 2000). This
6 contrasts with an increase in total yield (fruit weight) found in a previous study in California
7 (Takemotoetal., 1988).
8 When timothy was exposed in OTCs in Sweden to NF, CF, and CF+O3 treatments, there
9 was no effect of a 12-h mean O3 exposure of 68 ppb (NF treatment), but a 12-h mean exposure
10 of 152 ppb decreased yield by 58% (Danielsson et al., 1999). A similar lack of effect of
11 exposure to a 12-h mean O3 exposure of 62 ppb was found in a previous study in the United
12 States (Kohut et al., 1988).
13 Although most investigations of O3-response relationships focus on growth or yield of
14 marketable portions of plants, some studies also investigate effects on plant quality. In the study
15 of bahia grass in Alabama discussed above, in addition to the effects on yield, there were
16 significant effects on quality for ruminant nutrition (Muntifering et al., 2000). Concentrations of
17 neutral detergent fiber (NDF) were higher in primary-growth and regrowth forages from the
18 early-season planting when exposed to 2x-ambient O3 than when exposed to the NF treatment.
19 The concentration of acid detergent fiber was higher in the 2x-ambient treatment than in NF
20 treatment regrowth, whereas acid detergent lignin concentration was higher in 2 x-ambient than
21 in NF primary-growth forage. Crude protein concentrations were lower in CF-exposed than in
22 NF-exposed regrowth forage from the early planting and in CF- than in NF-exposed primary -
23 growth forage from the initial harvest of the late-season planting. No differences were observed
24 among treatments in concentrations of total phenolics in primary-growth or regrowth forages
25 from either planting, although concentrations of total phenolics tended to be higher in
26 CF-exposed than in NF-exposed primary-growth forage from the late-season planting. The
27 authors concluded that the alterations in quality of primary-growth and vegetative regrowth
28 forages were of sufficient magnitude to have nutritional and possibly economic implications to
29 their use for ruminant animal feed.
30 Sericea lespedeza and little bluestem were exposed to CF, NF, and 2 x-ambient O3 in OTCs
31 in Alabama for 10 weeks (Powell et al., 2003). Ozone treatments expressed as 12-h mean
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1 concentrations were 23, 40, and 83 ppb and expressed as seasonal SUM06 values were 0.2, 9.1,
2 and 61.0 ppm-h. Although there were few statistically significant effects of O3 on yield (the
3 yield of only the 2x-ambient compared to NF for Sericea lespedeza in the last of six harvests),
4 plant quality as feed for ruminants was reduced. The nutritive quality of Sericea lespedeza was
5 decreased by 7% and that of little bluestem by 2% as a result of increased cell wall constituents
6 and decreased in vitro digestibility.
7 For some annual species, particularly crops, the endpoint for an assessment of the risk of
8 O3 exposure can be defined as yield or growth; e.g., production of grain. For plants grown in
9 mixtures such as hayfields, and natural or seminatural grasslands (including native
10 nonagricultural species), endpoints other than production of biomass may be important. Such
11 endpoints include biodiversity or species composition and measures of plant quality such as total
12 protein, and effects may result from competitive interactions among plants in mixed-species
13 communities. Most of the available data on non-crop herbaceous species are for grasslands.
14 In a study of two perennial grasses (bent grass and red fescue) and two forbs (white clover,
15 Germander speedwell) grown in pots in OTCs, O3 effects differed among species and cutting
16 treatments (Ashmore and Ainsworth, 1995) (see also Table 9-18). Fescue biomass increased
17 with higher O3 treatments both in pots that were not cut during the growing season (mid-June to
18 mid-September), and those that were cut every two weeks. However, bent grass biomass
19 decreased with higher O3 exposure in the uncut treatment and increased in the cut treatment.
20 White clover and Germander speedwell biomass decreased substantially with higher O3 exposure
21 with and without cutting, with greater decreases in the cut treatment. The authors cautioned that
22 the experiment did not replicate field circumstances. The plants were all cut to 1 cm above the
23 ground, which does not simulate grazing, and there may have been effects due to growing the
24 plants in pots. However, two key results of this study likely apply to mixtures of species
25 growing in hay or forage fields or seminatural and natural communities. First, O3 exposure
26 increased the growth of O3-tolerant species while exacerbating the growth decrease of O3
27 sensitive species. Second, the total biomass of the mixed-species community was unaffected by
28 O3 exposure due to the differential effects on O3-sensitive and O3-tolerant species.
29 In a 2-year study using OTCs placed over managed pasture in Switzerland, the above-
30 ground biomass of clover (red and white) was reduced linearly in response to increased O3
31 exposure (Table 9-18) (Fuhrer et al., 1994). Exposure to a 12-h mean concentration of 39 ppb
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1 O3 reduced biomass by 24% as compared to the CF treatment with a 12-h mean concentration of
2 21 ppb O3. There was a trend toward increased above-ground biomass of grasses (primarily
3 orchard grass), but this trend was not statistically significant. As often found in other studies of
4 mixtures of species, by O3 exposure did not significantly affect total above-ground biomass O3
5 exposure.
6 A field-grown grass/clover mixture was exposed to CF, NF, and two O3 addition treatments
7 for two growing seasons in OTCs in southern Sweden (Pleijel et al., 1996). The mixture
8 consisted of 15% (by seed weight) red clover cv. Fanny, 60% timothy cv. Alexander, and 25%
9 fescue cv. Svalofs Sena.. Ozone concentrations expressed as AOT40 ranged from 0 to
10 approximately 47 ppm-h and expressed as 7-h mean from 11 to 62 ppb. Over this range, a slight,
11 but statistically significant, linear decrease of 4% in total above-ground biomass was seen
12 growth over six harvests. No significant decrease was seen in the proportion of clover, and the
13 authors ascribed this lack of effect to the relatively higher O3 sensitivity of timothy and lower
14 sensitivity of this clover cultivar as compared to previously published results for other
15 grass/clover mixtures (e.g., Fuhrer et al., 1994).
16 A mixture of species in an old farm field in Alabama was exposed to O3 for two growing
17 seasons in large OTCs 4.8 m high, and 4.5 m diameter); and a similar lack of effect of O3 was
18 found on total plant community growth measured as both canopy cover and vertical canopy
19 density (Barbo et al., 1998). Of the 40 species in the plots, O3 effects were examined only on the
20 most common species: blackberry, broomsedge bluestem, bahia grass, Panicum spp., and
21 winged sumac (second year only). Of these species, a 2x-ambient O3 treatment increased the
22 percent canopy cover of blackberry over 2 years by 124%, while that of winged sumac was
23 decreased by 95% (Table 9-18). Blackberry showed no significant effect on biomass, but ripe
24 fruit mass was decreased by 28% (Chappelka, 2002). However, there was a significant chamber
25 effect for this latter response. Biomass was not reported for other species in this study due to a
26 hurricane. Effects on loblolly pine grown in this experiment are discussed subsequently in
27 Section 9.6.5.5.
28 In summary, results of studies on perennial herbaceous species conducted since the 1996
29 criteria document was prepared are presented in Table 9-18. As for single-season agricultural
30 crops, yields of multiple-year forage crops are reduced at O3 exposures that occur in some years
31 over large areas of the United States (Table 9-14, Table 9-21). This result confirms that reported
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1 in the 1996 AQCD (U.S. Environmental Protection Agency, 1996). When species are grown in
2 mixtures, O3 exposure can increase the growth of O3-tolerant species while exacerbating the
3 growth decrease of O3-sensitive species (e.g., Ashmore and Ainsworth, 1995; Fuhrer et al.,
4 1994). Because of this competitive interaction, the total growth of the mixed-species community
5 may not be affected by O3 exposure (Ashmore and Ainsworth, 1995; Barbo et al., 1998; Fuhrer
6 et al., 1994). However, in some cases mixtures of grasses and clover species have shown
7 significant decreases in total biomass growth in response to O3 exposure in studies in the United
8 States (Heagle et al., 1989; Kohut et al., 1988) and in Sweden (Pleijel et al., 1996). In Europe,
9 a provisional critical level for perennials of an AOT40 value of 7 ppm-h over 6 months has been
10 proposed to protect sensitive plant species from the adverse effects of O3.
11
12 9.6.5.2 Deciduous Woody Species
13 It is extremely difficult and costly to study entire mature trees under controlled conditions
14 such as those in OTCs, with the possible exception of some species managed for fruit or nut
15 production. For this reason, the great maj ority of investigations have been of seedlings in
16 growth chambers, greenhouses, or OTCs. A few investigations have been carried out on
17 saplings or more mature trees using free air exposure systems (Haeberle et al., 1999; Isebrands
18 et al., 2000, 2001; Werner and Fabian, 2002). Exposure-response functions based on 28
19 experimental cases of seedling response to O3 suggest that a SUM06 exposure for 3 months of
20 31.5 ppm-h would protect hardwoods from a 10% growth loss in 50% of the cases (Table 9-19).
21 However, there is a substantial range in sensitivity among species. A risk analysis was
22 undertaken to predict tree biomass growth reductions due to O3 based on exposure-response
23 equations for seedlings of individual species combined with the species' spatial distribution
24 across the eastern United States and interpolated O3 exposure expressed as SUM06 (Hogsett
25 et al., 1997). The growth of sensitive species such as aspen and black cherry was predicted to be
26 reduced by at least 20% across 50% of their ranges in a high O3 year and approximately 10% in a
27 lower-than-average O3 year (Hogsett et al., 1997).
28 A few investigations reported since the last criteria document was prepared have examined
29 saplings or mature trees, notably of oak species in the southern Appalachian Mountains and pine
30 species in California. Most of these studies have been of natural (uncontrolled) O3 exposures.
31 Additional studies have examined foliar symptoms on mature trees, and in recent years such
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1 surveys have become more common and with greater attention to the standardization of methods
2 and the use of reliable indicator species (Campbell et al., 2000; Smith et al., 2003). Previous
3 criteria documents have noted the difficulty in relating foliar symptoms to effects on individual
4 tree growth, stand growth, or ecosystem characteristics (U.S. Environmental Protection Agency,
5 1996). This difficulty still remains to the present day.
6 Some investigators have suggested that a comprehensive risk assessment of the effects of
7 O3 on mature tree species might best be accomplished by extrapolating measured effects of O3
8 on seedlings to effects on forests using models based on tree physiology and forest stand
9 dynamics (Fuhrer et al., 1997; Hogsett et al., 1997; Chappelka, 1998; Laurence et al., 2000,
10 2001). Several such efforts are discussed in Sections 9.4 and 9.7.
11 In this subsection, emphasis will be placed on experimental evidence of O3 effects on the
12 growth of woody species under controlled conditions with some information from observational
13 studies under ambient conditions in forests. Experimental results are summarized for deciduous
14 species in Table 9-20; and the species are discussed below in the order in which they appear in
15 this table.
16 A series of studies out in Michigan and Wisconsin during the 1990s on clones of trembling
17 aspen previously demonstrated that they differ in their O3 sensitivity (Coleman et al., 1995b,
18 1995a, 1996; Dickson et al., 2001; Isebrands et al., 2000, 2001; Karnosky et al., 1996, 1998,
19 1999; King et al., 2001). Several of those studies were undertaken with plants in pots or in the
20 ground in OTCs and additional studies were undertaken at three sites selected to differ primarily
21 in O3 exposure (Karnosky et al., 1999). An ongoing study was undertaken using a free air
22 carbon dioxide and O3 enrichment (FACE) facility in Rhinelander, WI (Isebrands et al., 2000,
23 2001). These studies showed that O3-symptom expression was generally similar in OTCs and
24 FACE and gradient sites, supporting the previously observed variation among aspen clones
25 (Karnosky et al., 1999). In the Michigan OTC studies, O3 decreased total plant biomass by an
26 average of 18% in each of 2 years for three clones classified as high, intermediate and low in O3
27 tolerance (clones 216, 271 and 259; Karnosky et al. [1996]). However, in the first experiment,
28 an additional clone known to be O3 insensitive showed no response (clone 253). In these studies,
29 plants were grown in pots and exposed to CF, 1 x-ambient, and 2x -ambient O3 treatments in two
30 separate experiments of 98 days each (additional treatments of 0.5x-ambient and 1.5x-ambient
31 were used in the first experiment only (Karnosky et al., 1996). Ozone concentrations expressed
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1 as 3-month, 7-h mean values for 1990 ranged from 7 to 69 ppb for the CF to 2x-ambient
2 treatments; while for 1991, values ranged from 22 to 92 ppb.
3 The FACE study evaluated the effects of 3 years of exposure to combinations of elevated
4 CO2 and O3 on growth responses in mixture of five trembling aspen clones (Isebrands et al.,
5 2000, 2001). Height, diameter, and stem volume (diameter2 x height) were decreased by
6 elevated O3. On average for all clones, stem volume was decreased by 20% over the 3 years in
7 the elevated O3 treatment as compared with the 1 x-ambient treatment. However, one clone
8 showed increased growth in response to O3. Ozone concentrations were not reported. This
9 FACE facility study is important, because it confirmed responses reported previously with these
10 clones grown in pots or soil in OTCs, without the alterations of microclimate induced by
11 chambers. Currently, this is the only U.S. study using this technology to have examined the
12 effects of O3 under these conditions. This study is also significant, because the elevated
13 O3-exposure pattern used was intended to reproduce the 6-year average pattern from Washtenaw
14 County, Michigan (Karnosky et al., 1999).
15 Rooted cuttings of two aspen clones from Acadia National Park in Maine were exposed to
16 1 x-ambient, 1.7x-ambient, and 3x-ambient O3 concentrations in large OTCs in Ithaca, NY for
17 much of one growing season (15 June to 15 September) (Yun and Laurence, 1999a). Both
18 circular (4.7 m diameter, 3.7 m height) and rectangular (7.4 m x 2.75 m x 3.7 m height)
19 chambers were used (Mandl et al., 1989). Exposure to 1.7x-ambient O3 (SUM06 = 20 ppm-h,
20 9-h mean = 74 ppb) reduced shoot growth by 14 and 25% compared to ambient O3 for the two
21 clones (Yun and Laurence, 1999a). Total dry weight was reduced by 55 and 35% in the two
22 clones by the 3 x-ambient treatment (SUM06 = 62 ppm-h, 9-h mean =124 ppb) compared to the
23 ambient O3 treatment.
24 When black poplar cuttings in OTCs in Belgium were exposed to 8-h mean O3
25 concentrations of 5, 29, and 33 ppb, diameter growth decreased by 29% in the highest O3
26 treatment, but height growth was unaffected (Bortier et al., 2000b). A 2-month study of hybrid
27 poplar (Populus tremuloides x P. tremula) in a free air exposure system in Finland with 7-h
28 mean O3 concentrations of 30 and 38 ppb found a 6% decrease in height with no effect on
29 biomass (Oksanen et al., 2001). Eastern cottonwood cuttings in pots buried in the ground with
30 drip irrigation were exposed to ambient O3 at several sites in and near New York City in three
31 2-month experiments during three summers (Gregg et al., 2003). Ozone concentrations were
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1 lower at urban sites than at rural sites within 100 km of the urban sites. Total biomass growth
2 was greater in urban than rural sites, with a strong linear decrease in biomass with increasing O3
3 across all sites and years (r2 = 93). Total biomass decreased 33% with 12-h mean O3 levels of
4 38 ppb compared to 23 ppb. Multiple regression analysis showed no significant temperature
5 effect on biomass; therefore, the authors suggested that O3 exposures was the most likely
6 explanation for the reduced biomass in rural areas. The overall growth reductions and the
7 variation among genotypes seen on all of the above aspen and poplar studies is similar to those
8 previously reported in three OTCs studies summarized in the previous criteria document
9 (Table 9-26 of U.S. Environmental Protection Agency, 1996).
10 Black cherry seedlings grown in pots were exposed in OTCs in the Great Smoky Mountain
11 National Park in Tennessee to O3 treatments ranging from CF to 2x-ambient in two experiments
12 during 1989 and 1992 (Neufeld et al., 1995). Ozone exposure, expressed as SUM06, ranged
13 from 0 to 40.6 ppm-h in 1989, and from 0 to 53.7 ppm-h in 1992. Corresponding AOT40 values
14 were 0.0 to 28.3 ppm-h in 1989, and 0 to 40.4 ppm-h in 1992. In 1989, total biomass was
15 decreased in the 1.5x-ambient treatment by 18% and in the 2x-ambient treatment by 38%.
16 In 1992, total biomass was decreased in the 1.5 x-ambient treatment by 27%, and in the
17 2x-ambient treatment by 59%. In this study, SUM06 and AOT40 provided better fits than did
18 SUMOO with Weibull regression analyses to log-transformed biomass data. Although a Weibull
19 model was used, responses to O3 expressed as SUM06 and AOT40 appeared to be linear. The O3
20 exposures in the 1.5 x-ambient and 2 x-ambient treatments were reported to be similar to those
21 for a site near Charlotte, NC in a high-O3 year (1988). In a 2-year experiment in OTCs in Ohio,
22 seedlings of black cherry, sugar maple, and yellow poplar were exposed to O3 treatments with
23 SUMOO values ranging from 16 to 107 ppm-h in 1990 and 31 to 197 ppm-h in 1991 (Rebbeck,
24 1996). After two seasons of exposure, only black cherry showed a growth decrease: total
25 biomass was reduced by 32% in the 2x-ambient O3 treatment compared to the CF treatment; root
26 biomass was decreased by 39%. These results contrast with those of a previous study with black
27 cherry seedlings in which significant biomass reductions with exposures up to 2 x-ambient were
28 not observed (7-h mean = 21 to 97 ppb), perhaps because of the small sample size (3 seedlings
29 per chamber (Samuelson, 1994) in the earlier study.
30 A multiyear study of effects of O3 on both seedling and mature (30-year-old) red oak trees
31 was conducted in Norris, TN in large OTCs with three replicates per O3 treatment. Trees were
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1 exposed for 3 years to CF, 1 x-ambient, and 2x -ambient treatments, with the following
2 O3 exposures: SUM06 for 3 years = 0, 29, 326 ppm-h; SUMOO for 3 years = 147, 255, and
3 507 ppm-h. The net photosynthetic rate in mature trees was reduced by 25% in the ambient
4 treatment and by 50% in the 2x-ambient treatment (Hanson et al., 1994; Samuelson and
5 Edwards, 1993; Wullschleger et al., 1996). Despite these large decreases, no significant effects
6 on stem increment at the base, stem increment in the canopy, or leaf mass were observed for the
7 mature trees (Samuelson et al., 1996). Similarly, seedling biomass was not significantly reduced
8 by O3 exposure. The difficulty in replicating experiments with mature trees makes it difficult to
9 detect changes in growth or biomass. However, the mean values of the stem increment at the
10 base and within the canopy in the ambient treatment were larger than those in the CF treatment,
11 although those in the 2x-ambient treatment were lower. Therefore this study of mature trees
12 does not provide evidence that these ambient concentrations reduced aboveground tree growth,
13 even after 4-years exposure.
14 Sugar maple seedlings were exposed for 3 years to ambient, 1.7x-ambient, and 3 x-ambient
15 O3 treatments at both high-light (3 5% of ambient) and low-light levels (15% of ambient) (Topa
16 etal., 2001). This experiment was conducted in large OTCs near Ithaca, NY. Over the 3 years,
17 O3 exposures expressed as SUMOO for the three treatments were 88, 126, and 185 ppm-h. After
18 3 years, total seedling biomass in the 3 x-ambient treatment was reduced by 64% and 41% in the
19 low- and high-light treatments, respectively (compared to the 1 x-ambient treatment). The larger
20 reduction of biomass under low-light conditions suggests that seedlings growing under closed
21 canopies may be substantially more sensitive to O3 than are seedlings exposed to higher-light
22 levels in gaps or clearings. These results differ from other studies in which seedling biomass
23 was unaffected by exposure to SUMOO values of 304 ppm-h over 2 years (Rebbeck, 1996) or
24 591 ppm-h over 3 years (daytime mean of 40.7) (Laurence et al., 1996). However the latter two
25 studies used much higher light levels, which may have reduced the O3 effect, based on the results
26 of Topa etal. (2001).
27 Although most studies demonstrate that O3 decreases biomass growth, occasional results
28 indicate that O3 can increase growth of some portions of woody perennials. When Casselman
29 plum trees near Fresno, CA were exposed to O3 in large, rectangular OTCs to three O3 treatments
30 (CH, 1 x-ambient, and an above-ambient O3 treatment) for 4 years (12-h mean = 31, 48, 91
31 ppm-h), stem increment increased 14% in the highest O3 treatment compared to the CH
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1 treatment; and this effect was statistically significant (Retzlaff et al., 1997). However, fruit yield
2 decreased in this treatment by 42% and also decreased by 16% in the 1 x-ambient-O3 treatment.
3 Root growth was not measured in this study. Hence, the increase in stem diameter may have
4 been at the expense of other organs. However, in a fifth year, all plants were exposed to
5 1 x-ambient O3, and there were no differences in fruit yield, suggesting that trees were able to
6 recover to some extent from the effects of O3 exposure in prior years.
7 When yellow poplar seedlings were exposed to O3 concentrations up to SUMOO values of
8 107 ppm-h in one year and 197 ppm-h in a second year, no effects on biomass were observed
9 (Rebbeck, 1996).
10 Many studies have demonstrated that root growth is more sensitive to O3 exposure than is
11 stem growth. For example, in a study with black cherry seedlings exposed in OTCs in
12 Tennessee in 1989, root biomass in a 2 x-ambient treatment was decreased by 42%, while stem
13 biomass was decreased by only 24%. However, in a second experiment in 1992, root and stem
14 growth reductions in the 2x-ambient treatment were similar (65% versus 60%) (Neufeld et al.,
15 1995). In Finland, reduced root growth was found for a number of clones of silver birch
16 (Oksanen and Saleem, 1999). After 5 years, root growth was decreased by 33%, but shoot
17 growth was not affected by O3 exposures of a 7-h mean of 15 ppm-h over 5 years in a FACE
18 system (Oksanen, 2001). When first-year poplar seedlings (P. tremuloides) were exposed in
19 OTCs to two O3 concentrations and six N concentrations, the root/shoot ratio was decreased soon
20 after exposure to O3 in all N treatments, even though O3 effects on total biomass were not
21 detected in the low-N and very high-N treatments (Pell et al., 1995). These results suggest that
22 effects on the root/shoot ratio occur before significant growth effects arise. In a series of OTC
23 experiments lasting 1 to 3 years at 3 different elevations in Switzerland, fine root growth in
24 European beech was found to be more sensitive to O3 than was shoot or total biomass (Braun and
25 Fluckiger, 1995). Although the estimated effect of O3 on fine root biomass was similar to that
26 for total biomass, fine root biomass was significantly decreased at AOT40 (24-h) values of only
27 3 ppm-h, while total biomass was not significantly decreased until AOT40 values reached 30 to
28 40 ppm-h.
29
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1 9.6.5.3 European Critical Levels
2 In Europe, a Level I critical level has been set for forest trees based on OTC studies of
3 saplings. This level is discussed here because it was based on a deciduous tree species.
4 For consistency with the approach used for crops, an AOT40 index value was selected. A few
5 studies have shown that O3 can be taken up by tree species at nighttime, e.g., young birch trees
6 (Matyssek et al., 1995). However, because most evidence suggests that O3 deposition at
7 nighttime is low (Coe et al., 1995; Rondon et al., 1993), a value for only daylight hours was
8 selected in Europe (Fuhrer et al., 1997; Karenlampi and Skarby, 1996). European beech was
9 selected for development of a Level I critical level, because data from several studies were
10 available for this species and because deciduous tree species were judged to be more sensitive to
11 O3 compared to evergreen tree species (Fuhrer et al., 1997; Karenlampi and Skarby, 1996).
12 A critical level was defined as an AOT40 value of 10 ppm-h for daylight hours for a 6-month
13 growing season (Karenlampi and Skarby, 1996). However, other studies have shown that other
14 species such as silver birch may be more sensitive to O3 than beech (Paakkonen et al., 1996). As
15 discussed for annual plants above (Section 9.6.4.6), Level I critical values are not designed for
16 making quantitative estimates of the O3 effects on vegetation at the regional scale. For long-
17 lived perennials, additional problems complicate extrapolation. As discussed below (Section
18 9.6.5.7), considerable scaling is required to extrapolate from experiments conducted with tree
19 seedlings to estimate effects on mature trees in forests. Because of these scaling issues, there is
20 greater uncertainty in estimating effects on forest trees than on annual plants such as crops.
21 While some information is available for addressing issues such as scaling from seedlings to
22 mature trees and estimating O3 uptake, this information may still be insufficient for developing a
23 Level II approach that can provide quantitative estimates of forest growth losses due to O3
24 exposure (Broadmeadow, 1998).
25
26 9.6.5.4 Summary of Effects on Deciduous Woody Species
27 Recent evidence from free air exposure systems and OTCs supports results observed
28 previously in OTC studies (Table 9-16, Figure 9-1). Specifically, a series of studies undertaken
29 using free air O3 enrichment in Rhinelander, WI (Isebrands et al., 2000, 2001) showed that
30 O3-symptom expression was generally similar in OTCs, FACE, and ambient-O3 gradient sites,
31 supporting the previously observed variation among aspen clones using OTCs (Karnosky et al.,
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1 1999). New evidence is also available comparing various aspects of O3 sensitivity between
2 seedlings and mature trees of some species, notably red oak. As has been observed in previous
3 criteria documents, root growth is often found to be the most sensitive indicator in terms of
4 biomass response to O3.
5 Results since 1996 support the conclusions of the 1996 AQCD (U.S. Environmental
6 Protection Agency, 1996) that individual deciduous trees are generally less sensitive to O3 than
7 are most annual plants, with the exception of a few genera such as Populus, which are highly
8 sensitive. However, the data presented in Table 9-19 suggest that ambient exposures that occur
9 in different regions of the United States can sometimes reduce the growth of seedlings of
10 deciduous species. Results from multiple-year studies sometimes show a pattern of increasing
11 effects in subsequent years. Although, in some cases, growth decreases due to O3 become less
12 significant or even disappear over time. While some mature trees show greater O3 sensitivity in
13 physiological parameters such as net photosynthetic rate compared to seedlings, these effects
14 may not translate into measurable reductions in biomass growth. Because even multiple-year
15 experiments do not expose trees to O3 for more than a small fraction of their life span and
16 because competition may, in some cases, exacerbate the effects of O3 on individual species,
17 determining the effects on mature trees remains a significant challenge. Effects on mature trees
18 under natural conditions are discussed after the review of evergreen species below and more
19 fully in Section 9.7, in the context of extrapolating from controlled studies to forest ecosystems.
20
21 9.6.5.5 Evergreen Woody Species
22 Most investigations have shown evergreen tree species to be less sensitive to O3 compared
23 to deciduous species (U.S. Environmental Protection Agency, 1996). For example, exposure-
24 response functions based on 23 experimental cases of seedling response to O3, suggest that a
25 SUM06 exposure for 3 months of 42.6 ppm-h would protect evergreen species from a 10%
26 growth loss in 50% of the cases (Table 9-16). For deciduous species, the corresponding SUM06
27 value was 31.5 ppm-h (Table 9-16). As another example, experiments in the Great Smoky
28 Mountains National Park found black cherry seedlings to demonstrate substantial decreases in
29 biomass, as discussed above and shown in Table 9-19 (Neufeld et al., 1995). However, exposure
30 for up to three growing seasons did not decrease the biomass of eastern hemlock, Table
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1 Mountain pine, and Virginia pine seedlings exposed to O3 under similar conditions in this
2 location, as shown in Table 9-19 (Neufeld et al., 2000).
3 As for deciduous species, there is a substantial range in sensitivity among evergreen
4 species. As discussed above for deciduous species, a risk analysis was undertaken to predict tree
5 biomass growth reductions due to O3 based on exposure-response equations for tree seedlings
6 combined with species distribution across the eastern United States and interpolated O3 exposure
7 (Hogsett et al., 1997). While some species such as Virginia pine were predicted to be affected
8 only slightly even in a high O3 year, the growth of sensitive evergreen species such white pine
9 was predicted to be reduced by 5% in a lower-than-average O3 year and 10% in a high ozone
10 year across 50% of its range (Andersen et al., 1997). The remainder of this section discusses
11 experimental results for evergreen species in the order shown in Table 9-20.
12 Douglas fir seedlings were exposed to elevated O3 concentrations in a free air zonal air
13 pollution system in British Columbia, Canada for two growing seasons with 12-h mean values in
14 1988 of 18 to 41 ppb and in 1989 of 27 to 66 ppb (Runeckles and Wright, 1996). Although
15 substantial variation was seen in effects among the different treatments, there was a significant
16 decrease in the growth of the second flush weight in the second year, with reductions of 55% at
17 the highest O3 exposure, based on a linear regression. This result contrasts with the lack of
18 effect seen in a previous study with seedlings of this species grown in pots for 134 days and
19 exposed to 7-h mean O3 concentrations up to 71 ppb (Table 9-30 in U.S. Environmental
20 Protection Agency, 1996).
21 First-year loblolly pine seedlings of 53 open-pollinated families were exposed to
22 1 x-ambient O3 in OTCs for a single growing season, and average growth volume was decreased
23 by 14% compared to a CF treatment (McLaughlin et al., 1994). The 1 x-ambient O3 exposure in
24 this study expressed as 24-h SUMOO was 137 ppm-h, and the CF treatment reduced O3 by 47%.
25 In this study, the root/shoot ratio was decreased significantly in six of the nine families
26 examined. Exposure to O3 with SUM06 values up to 162 ppm-h and 132 ppm-h in 2 successive
27 years in OTCs had no effect on seedlings grown in competition with various species of grasses
28 and forbs (Barbo et al., 2002). Exposure of 3-year-old seedlings to O3 exposures of up to
29 2.5x-ambient (12-h mean of 98 ppb) also had no significant effect (Anttonen et al., 1996).
30 Information summarized in the 1996 AQCD (U.S. Environmental Protection Agency, 1996),
31 indicated that significant effects on seedling growth were observed in several studies of
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1 seedlings exposed to elevated O3 concentrations for one or more years. Several studies,
2 including that of McLaughlin et al. (1994), demonstrate considerable variation in O3 sensitivity
3 among different genotypes of loblolly pine.
4 For ponderosa pine seedlings, the 1996 AQCD reviewed a number of studies with
5 exposures to elevated O3 concentrations for one to three growing seasons (U.S. Environmental
6 Protection Agency, 1996). Recent similar studies support the earlier results (Table 9-20);
7 (Andersen et al., 2001; Takemoto et al., 1997). The 1996 criteria document also discussed at
8 some length the ongoing work examining effects of O3 across a naturally-occurring O3 gradient
9 in the San Bernardino Mountains in California. Since that time, much research on ponderosa
10 pine has focused on interactive effects of additional stresses such as nitrogen, and effects of O3
11 on physiological parameters (Sections 9.4, 9.7). Effort has also focused on the effects of O3 on
12 root growth because such effects could alter sensitivity to drought or nutrient stress. Ecosystem
13 level effects of ozone are discussed further in Section 9.7, but some information relevant to
14 exposure-response relationships is discussed below.
15 For several tree species, O3 has been shown in experimental studies with seedlings to
16 reduce root growth more than shoot growth (U.S. Environmental Protection Agency, 1996).
17 Ponderosa pine has been shown to be sensitive to O3, and studies with seedlings have shown
18 reduced root growth, decreases root/shoot ratios, and decreased allocation to roots (Andersen
19 et al., 1991; 1997; Andersen and Rygiewicz, 1995; Andersen and Scagel, 1997; Tingey et al.,
20 1976). More recently, data from a long-studied pollution gradient in the San Bernardino
21 Mountains of southern California suggests that O3 substantially reduces root growth in natural
22 stands of ponderosa pine. Root growth in mature trees was decreased at least 87% in a high
23 pollution site as compared to a low pollution site (Grulke et al., 1998), and a similar pattern was
24 found in a separate study with whole tree harvest along this gradient (Grulke and Balduman,
25 1999). Because other potential influences on root growth, including shading by competing trees,
26 soil temperature, soil moisture, phenology, were not correlated with the observed pattern of
27 reduced root growth, the authors conclude that O3 was the cause of the observed decline in root
28 growth. Further results of field investigations with ponderosa pine and other pine species native
29 to California are discussed below under the heading "Scaling experimental data on long-lived
30 species to field conditions" as well as in Section 9.7.
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1 Table Mountain pine, Virginia pine, and eastern hemlock seedlings were exposed to
2 various levels of O3 (CF to 2 x ambient) in OTCs for in a series of experiments two or three
3 years in Great Smoky Mountains National Park in Tennessee (Neufeld et al., 2000). There were
4 no statistically significant effects of O3 exposure on stem or root biomass, and only slight effects
5 on the biomass of the oldest needles in Table Mountain pine in the 2* ambient treatment.
6 As reviewed in the 1996 criteria document, studies of the response of red spruce to
7 O3 exposures generally have not found effects on growth of seedlings or saplings, even after
8 exposure to high concentrations (12-h mean of 90 ppb) for up to 4 years. A more recent report
9 confirms that this slow-growing species is O3 insensitive for at least several years (Laurence
10 etal., 1997).
11 For perennial vegetation, cumulative effects over more than one growing season may be
12 important. For three-year-old Norway spruce in Sweden, exposure to elevated O3 for three
13 growing seasons decreased total biomass by 18% and stem biomass by 28% (Karlsson et al.,
14 1995). However, after a fourth season of exposure, total biomass decreased significantly by only
15 8% (Karlsson et al., 2002). In this experiment, the O3 exposures expressed as 12-h mean values
16 averaged over four growing seasons were 12 and 44 ppb for the CF and 1.5 x-ambient treatments,
17 respectively; and AOT40 values were 2 and 23 ppm-h, respectively. Despite 4 years of
18 exposure, this experiment did not demonstrate a consistent trend in the O3 effect on biomass to
19 indicate a significant carry-over effect.
20
21 9.6.5.6 Summary of Effects on Evergreen Woody Species
22 In summary, the O3 sensitivity of different genotypes within species and between species
23 of evergreen vegetation varies widely. Based on studies with seedlings in OTCs, major species
24 in the United States are generally less sensitive than are most deciduous trees, and slower-
25 growing species are less sensitive than faster-growing ones. However, exposure to ambient O3
26 may reduce the growth of seedlings of commonly occurring species. Because tree species are
27 long-lived, most experiments have only covered a very small portion of the life span of a tree,
28 making estimating of any effect on mature trees difficult. Considerations for scaling the results
29 of seedling studies to mature forest trees as well as additional information from field surveys and
30 studies of mature trees under natural conditions are discussed below and in Section 9.7.
31
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1 9.6.5.7 Scaling Experimental Data to Mature Trees
2 As compared with annual crop species, it is much more difficult to define appropriate
3 exposure-response relationships for tree species. For annual species, an experiment may cover
4 the whole life span of the plant, but it is not feasible to provide controlled-exposure conditions
5 for long-lived plants for any significant portion of their life spans. Most studies have used small
6 seedlings because they are manageable under experimental conditions; but seedlings and mature
7 trees may have different sensitivities to O3. For perennial species, effects of O3 may accumulate
8 over more than 1 year, and may interact with other stresses such as drought stress over multiple
9 growing seasons. As for annual species (Section 9.4.2), substantial variability occurs among
10 evergreen genotypes and this variation may interact with other stress responses differently in
11 different landscapes and regions. Despite these difficulties, investigators have addressed some
12 of these issues since the 1996 AQCD (U.S. Environmental Protection Agency, 1996). New
13 information is available on the response of mature evergreen trees to O3 under field conditions,
14 and models based on tree physiology and stand dynamics have been used to predict O3 effects on
15 forest stands and regions. The following issues are reviewed briefly below: (1) interaction of
16 drought and O3 stress, (2) scaling data from seedlings to mature-tree studies. Two additional
17 scaling issues are addressed in Section 9.7: (1) scaling data to forest stands, and (2) scaling data
18 to ecosystems and regions.
19
20 9.6.5.7.1 Interactive Effects of Drought and Ozone
21 Many interacting factors may influence the effect of O3 on vegetation. For crop plants,
22 environmental conditions are often managed such that nutrients and water are not strongly
23 limiting; but for native vegetation, including most perennial species, such factors are likely to
24 limit growth. The effects of interacting stresses on vegetation are reviewed in Section 9.4.
25 However, because drought is common in many forests, and because there is substantial evidence
26 that it alters the response of trees to O3, it is discussed in this section in the context of
27 determining exposure-response relationships for trees.
28 Controlled experiments with seedlings provide direct evidence that drought can reduce the
29 impact of O3. For example, for 3-year-old Norway spruce in Sweden, exposure to elevated
30 O3 for three growing seasons decreased total biomass by 18% and stem biomass by 28%
31 (Karlsson et al., 1995). However, for draughted trees, both total and stem biomass decreased
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1 only 5%, with a statistically significant interaction with O3 for stem biomass. Yet after a fourth
2 season of exposure, there was no longer any interaction between drought and O3, while there was
3 a significant decrease of 8% in the biomass when both drought and well-watered data were
4 combined (Karlsson et al., 2002). In this study, seedlings were grown in sand in 120 L pots and
5 for the drought treatment, water was withheld for 4 weeks during the first year and for 7 to
6 8 weeks during each of the last 3 years. In this experiment, the O3 exposures expressed as 12 h
7 seasonal daylight mean averaged over four growing seasons were 12 and 44 ppb for the CF and
8 1.5x-ambient treatments respectively. Over this period, the AOT40 values for the treatments
9 averaged 2 and 23 ppm-h respectively. Despite 4 years of exposures, this experiment did not
10 demonstrate a consistent trend in drought O3 interactions. The difference in effects seen between
11 the third and fourth season suggest that scaling drought-O3 interactions from seedlings to mature
12 trees may be difficult. However, evidence from biomonitoring surveys supports an interaction
13 between drought and O3 effects, at least for foliar symptoms. In systematic surveys of foliar
14 symptoms on species selected as biomonitors throughout much of the eastern United States,
15 symptoms were more common and more severe in areas with high O3 concentrations (Smith
16 et al., 2003). However, in very dry years, such as 1999, the occurrence and severity of
17 symptoms was greatly reduced, even in areas with high ambient O3 concentrations.
18
19 9.6.5.7.2 Scaling from Seedlings to Mature Trees
20 Because most experiments are conducted with seedlings, various methods are required to
21 scale experimental data on seedlings to mature trees. An overview of physiological differences
22 between young and old plants, and the consequences of these differences for O3 sensitivity, is
23 provided in Section 9.4.5.3. The discussion below focuses on information relevant to developing
24 exposure-response relationships for mature trees. Information from a few experimental studies,
25 as well as scaling efforts based on physiological characteristics incorporated into models, are
26 discussed in Section 9.7.
27 Although most studies continue to examine the effects of O3 on seedlings, during the 1990s
28 some studies examined the effects of O3 on the response of mature trees. Studies of mature trees
29 demonstrate differences in some aspects of O3 sensitivity between seedlings and mature trees.
30 For some species, such as red oak, seedlings are less sensitive to O3 than are mature trees
31 (Hanson et al., 1994; Samuelson and Edwards, 1993; Wullschleger et al., 1996). Both red oak
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1 seedlings and genetically related mature trees were exposed to CF, 1 x -ambient, or 2 x-ambient
2 O3 exposures in OTCs in Tennessee for two growing seasons (Hanson et al., 1994). Nine large
3 chambers (4.6 x 8.2 m) were used to enclose individual mature trees and standard EPA-style
4 OTCs were used for potted seedlings. Ozone exposures expressed as a 24-h SUMOO were 34,
5 79, and 147 ppm-h in 1992 and 37, 95, 188 ppm-h in 1993 for the sub-ambient, and 2x-ambient
6 treatments. Mature trees had a greater light-saturated net photo synthetic rate and stomatal
7 conductance compared to seedling foliage at physiological maturity. By the end of the growing
8 season, exposure to 1 x-ambient and 1 x-ambient O3 reduced the light-saturated net
9 photosynthetic rate and stomatal conductance of mature trees by 25 and 50%, respectively,
10 compared with the CF treatment (35 ppm-h). In seedlings, however, light-saturated net
11 photosynthetic rate and stomatal conductance were less affected by O3 exposure. The authors
12 concluded that extrapolations of the results of seedling-exposure studies to foliar responses of
13 mature forests without considering differences in foliar anatomy and stomatal response between
14 juvenile and mature foliage may introduce large errors into projections of the O3 responses of
15 mature trees.
16 In a study of ponderosa pine in California, seedlings and branches of mature trees
17 (in branch chambers) were exposed to O3 concentrations of 0.5-, 1-, and 2x-ambient O3
18 concentrations (Momen et al., 1997). Net photosynthetic rate of 1-year-old, but not current-year,
19 foliage was reduced in mature trees while not significantly reduced in seedlings. This effect was
20 not due to alteration of stomatal conductance by O3. This result contrasts with those with earlier
21 studies of red spruce (Rebbeck et al., 1993).
22 In contrast to the findings for red oak and ponderosa pine, giant sequoia seedlings had
23 higher rates of stomatal conductance, CO2-exchange rate, and dark respiration than did mature
24 trees (Grulke and Miller, 1994). As compared to older trees, stomatal conductance was more
25 than 7-fold greater in current-year and 4-fold greater in 2-year-old seedlings (Grulke and Miller,
26 1994). The authors concluded that giant sequoia seedlings are sensitive to atmospheric O3 until
27 ~5 years of age. Low conductance, high water use efficiency, and compact mesophyll all
28 contribute to a natural O3 tolerance, or to O3 defense, or to both, in the foliage of older trees.
29 Similarly, lower stomatal conductance was found in mature Norway spruce in Austria compared
30 to seedlings grown with optimal water and nutrients in a growth chamber (Wieser, 1997). In this
31 study, net photosynthetic rate was less sensitive to added O3 in mature trees compared to
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1 seedlings. In a related study, the average rate of O3 uptake of 17-year-old trees
2 -0.6 nmol nT2 s"1, decreasing linearly in older trees, such that rates were only ~0.1 nT2 s"1 in
3 216-year-old trees (Wieser et al., 1999).
4 Based on a review of studies of stomatal conductance in both seedlings and mature trees,
5 Samuelson and Kelly (2001) concluded that O3 uptake in oak species, black cherry, sugar maple,
6 and American beech averaged 47% lower in potted seedlings than in mature trees. For evergreen
7 species, they concluded that O3 uptake in seedlings averaged 26% higher than in mature trees.
8 They also suggested that artifacts introduced by growth in pots confound these differences that
9 exposure-response functions derived from seedlings grown in situ are more applicable to mature
10 trees than are studies of seedlings grown in pots (Samuelson and Kelly, 2001).
11 As discussed above for annual vegetation, it has long been noted that internal O3 dose is
12 more appropriate than external O3 exposure for assessing the effects of O3 on vegetation, because
13 effects occur primarily via the uptake of O3 through the stomata (Section 9.3.2). However,
14 external O3 exposure sometimes has been shown to explain O3 effects as well or better than
15 calculated internal O3 dose. For ponderosa pine, Grulke and others (2002) found little difference
16 in the response of net photosynthetic rate and stomatal conductance to O3 exposure as compared
17 to calculated O3 uptake; and estimated O3 uptake by ponderosa pine and O3 exposure at several
18 sites were highly correlated (r2 = 0.92). For red oak, Hanson and others (1994) found that
19 SUMOO explained 83% of the variance in the response of light-saturated photosynthetic rate to
20 O3 levels, while estimated internal dose explained only 76% of the variance. In this same study,
21 SUM06 explained only 49% of the variance. Due to genetic variation or other factors, individual
22 mature trees will vary in their response to similar O3 exposures. For example, in 125-year-old
23 giant sequoia trees exposed to -230 ppm-h of O3 in branch chambers, O3 uptake in one
24 individual was -5 mmol nT2, while in another it was -9.5 mmol nT2 (Grulke et al., 1996).
25 Based on these results, stomatal conductance, O3 uptake, and O3 effects cannot be assumed
26 to be equivalent in seedlings and mature trees. In general, mature deciduous trees are likely to
27 be more sensitive to O3 compared to seedlings, while mature evergreen trees are likely to be less
28 sensitive than seedlings. However, even when differences in physiological traits occur,
29 concomitant effects on stem growth may not be detected in the field. Additionally, complex
30 interactions may occur between environmental conditions and O3 responses;and artifacts may
31 occur for seedling studies, especially for seedlings grown in pots. Section 9.7 further discusses
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1 issues that must be addressed when scaling data from individual mature trees to forests and
2 regions.
3
4 9.6.6 Studies with the Chemical EDU
5 The chemical EDU (7V-[2-(2-oxo-l-imidagolidinyl)ethyl]-7V-phenylurea) has been used
6 with the goal of protecting plants from O3 effects without controlling O3 exposure (Table 9-21)
7 (U.S. Environmental Protection Agency, 1986, 1996). The use of EDU has the potential to be a
8 low-cost, practical method of evaluating ambient O3 exposures on plants grown under natural
9 conditions without the limitations imposed by methodologies such as OTCs (Section 9.2). EDU
10 can be applied easily as a foliar spray or a soil drench and, with more effort, can be injected
11 directly into plant stems. However, because EDU is phytotoxic, and may have effects on plants
12 other than antioxidant protection, it is crucial that the correct dosage for protection from O3, be
13 determined without the direct effects of EDU. Unfortunately, although many studies with EDU
14 have been conducted in recent decades, very few have used multiple EDU application levels
15 along with multiple O3 exposures to characterize the EDU system for a given plant species.
16 Therefore, the text of this section focuses on how data from existing studies can be used for
17 developing or validating exposure-response relationships, rather than reviewing results of all
18 individual studies. Data from individual studies on O3 exposure, EDU application rates, and the
19 effects of EDU are presented in Table 9-21. In addition to EDU, sodium erythorbate has been
20 used in a few studies as a protectant chemical. Since very few published studies have used
21 sodium erythorbate and attempts to establish appropriate doses for individual species are even
22 more limited, the use of this chemical is not reviewed here.
23 The phytotoxicity of EDU is well known, and the point has been made repeatedly that for
24 a particular species or cultivar, tests under a range of environmental conditions and O3 exposures
25 must be made to establish the efficacy of EDU for quantifying O3 effects (Heggestad, 1988;
26 Kostka-Rick and Manning, 1992). A recent study has shown that even low concentrations of
27 EDU (8 to 32 mg I/1 soil), decreased bean yield under low O3 exposure (7-h mean of 19 ppb) in
28 CF OTCs (Miller et al., 1994). This study also demonstrated that phytotoxicity (both foliar
29 symptoms and growth effects) can differ even in the same series of experiments, apparently due
30 to changes in environmental conditions, and that EDU can suppress yield at application rates that
31 do not always cause foliar symptoms (Miller et al., 1994). Finally, this study found that EDU
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1 altered biomass partitioning by increasing vegetative growth and decreasing reproductive
2 growth. A study of bean grown in OTCs in Germany found that EDU treatment in CF OTCs
3 significantly increased yield, while EDU had no significant effect on yield in other O3 treatments
4 (Brunschon-Harti et al., 1995). In this study, O3 significantly reduced the mass of pods, shoots,
5 and roots. EDU increased root, leaf, and shoot mass across O3 treatments. However, there was
6 only a significant interaction with O3 for root mass. This study indicates that EDU can stimulate
7 above-ground growth and/or delay senescence regardless of O3 treatment. Together, these
8 studies suggest that EDU has effects other than its antioxidant protection and phytotoxicity.
9 Recent studies have also shown that EDU does not always have greater effects at higher O3
10 exposures. For bean grown in pots in studies in Spain and the Netherlands, EDU increased pod
11 yield (Ribas and Penuelas, 2000; Tonneijck and Van Dijk, 1997). However, this effect was not
12 greater at sites with higher O3 exposure despite consistent experimental protocols at all sites,
13 including growing the same cultivar in pots with adequate water (Ribas and Penuelas, 2000;
14 Tonneijck and Van Dijk, 1997). Such results suggest that it may be difficult to quantify
15 ambient O3 effects using EDU, because the amount of plant growth or yield expected at a low
16 (background) O3 concentration cannot be inferred from EDU-treated plants grown at locations
17 with higher O3 exposures.
18 Potted white clover exposed to ambient O3 at 12 sites throughout Europe over
19 three growing seasons was evaluated in a meta-analysis (Ball et al., 1998). A soil drench of
20 100 mL EDU was applied every 2 weeks for 3 months per growing season. Only very weak
21 evidence of a linear relationship of the ratio of biomass in control versus EDU treated plants
22 across all sites (r2 = 0.16) was seen. However, an artificial neural network (ANN) model
23 including VPD, temperature, longitude, year, and altitude explained much more of the variance
24 (r2 = 0.79). The authors suggested that the greater sensitivity at certain sites in Germany may
25 have been due to occurrence of other pollutants. This meta-analysis indicates that EDU effects
26 may be influenced substantially by environmental factors.
27 In summary, the EDU method for assessing the impact of ambient O3 exposures is
28 potentially useful, because it provides a separate line of evidence than other methods. However,
29 as has been pointed out by numerous authors, the system must be carefully characterized for
30 each species under different environmental conditions and different O3 exposures.
31 Unfortunately, such characterization has so far been limited, although substantial progress has
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1 been made for radish (Kostka-Rick et al., 1993; Kostka-Rick and Manning, 1992, 1993). Effects
2 of EDU can include phytotoxicity and alteration of biomass partitioning, along with differential
3 effects on the same species in different years or locations. Additionally, the degree of O3
4 protection afforded by a certain application of EDU is difficult to quantify. Thus, it is difficult to
5 use data from existing EDU studies to develop exposure-response relationships or to quantify the
6 effects of ambient O3 exposure. Despite these limitations, the EDU studies reviewed in previous
7 criteria documents (U.S. Environmental Protection Agency, 1986, 1996) and the more recent
8 studies summarized in Table 9-21 provide another line of evidence that ambient O3 exposures
9 occurring in many regions of the United States may be reducing the growth of crops and trees.
10
11 9.6.7 Summary
12 Recently published data support the conclusions of previous criteria documents that there
13 is strong evidence that ambient concentrations of O3 cause injury and damage to numerous
14 common and economically valuable plant species. For annual vegetation, the data summarized
15 in Table 9-17 show a range of growth and yield responses both within and among species.
16 Nearly all of these data were derived from studies in OTCs, with only two studies using open-air
17 systems in the UK (Ollerenshaw et al., 1999; Ollerenshaw and Lyons, 1999). It is difficult to
18 compare studies that report O3 exposure using different indices, such as AOT40, SUM06, or 7-h
19 or 12-h mean values. However, when such comparisons can be made, the results of recent
20 research confirm earlier results summarized in the 1996 AQCD (U.S. Environmental Protection
21 Agency, 1996). A summary of earlier literature concluded that a 7-h, 3-month mean of 49 ppb
22 corresponding to a SUM06 exposure of 26 ppm-h would cause 10% loss in 50% of 49
23 experimental cases (Tingey et al., 1991). Recent data summarized in Table 9-17 support this
24 conclusion, and more generally indicate that ambient O3 exposures can reduce the growth and
25 yield of annual species. Some annual species such as soybean are more sensitive, and greater
26 losses may be expected (Table 9-17). Thus the recent scientific literature supports the
27 conclusions of the 1996 AQCD (U.S. Environmental Protection Agency, 1996) that ambient O3
28 concentrations are probably reducing the yield of major crops in the United States.
29 Much research in Europe has used the AOT40 exposure statistic, and substantial effort has
30 gone into developing Level-1 critical levels for vegetation using this index. Based on regression
31 analysis of 15 OTC studies of spring wheat including one study from the United States and
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1 14 from locations ranging from southern Sweden to Switzerland, an AOT40 value of 5.7 ppm-h
2 was found to correspond to a 10% yield loss, and a value of 2.8 ppm-h corresponded to a 5%
3 yield loss (Fuhrer et al., 1997). Because a 4 to 5% decrease could be detected with a confidence
4 level of 99%, a critical level of an AOT40 value of 3 ppm-h was selected in 1996 (Karenlampi
5 and Skarby, 1996).
6 In addition to likely reductions in crop yield, O3 may also reduce the quality or nutritive
7 value of annual species. Many studies have shown effects of O3 on various measures of plant
8 organs that affect quality, with most studies focusing on characteristics important for food or
9 fodder. These studies indicate that there may be economically important effects of ambient O3
10 on the quality of crop and forage species. Previous criteria documents have concluded that
11 visible symptoms on marketable portions of crops and ornamental plants can occur with seasonal
12 7-h mean O3 exposures of 40 to 100 ppb (U.S. Environmental Protection Agency, 1978, 1986,
13 1996). The recent scientific literature does not refute this conclusion.
14 The use of OTCs may reverse the usual vertical gradient in O3 that occurs within a few
15 meters above the ground surface (Section 9.2). This reversal suggests that OTC studies may
16 somewhat overestimate the effects of an O3 concentration measured several meters above the
17 ground. However such considerations do not invalidate the conclusion of the 1996 AQCD (U.S.
18 Environmental Protection Agency, 1996) that ambient O3 exposures (Tables 9-14 and
19 Table 9-22) are sufficient to reduce the yield of major crops in the United States.
20 As for single-season agricultural crops, yields of multiple-year forage crops are reduced at
21 O3 exposures that occur over large areas of the United States. This result is similar to that
22 reported in the 1996 AQCD (U.S. Environmental Protection Agency, 1996). When species are
23 grown in mixtures, O3 exposure can increase the growth of O3- tolerant species and exacerbate
24 the growth decrease of O3-sensitive species (e.g., Ashmore and Ainsworth, 1995; Fuhrer et al.,
25 1994). Because of this competitive interaction, the total growth of the mixed-species community
26 may not be affected by O3 exposure (Ashmore and Ainsworth, 1995; Barbo et al., 1998; Fuhrer
27 et al., 1994). However, in some cases, mixtures of grasses and clover species have shown
28 significant decreases in total biomass growth in response to O3 exposure in studies in the United
29 States (Heagle et al., 1989; Kohut et al., 1988) and in Sweden (Pleijel et al., 1996). In Europe,
30 a provisional critical level for herbaceous perennials of an AOT40 value of 7 ppm-h over
31 6 months has been proposed to protect sensitive plant species from adverse effects of O3.
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1 For deciduous tree species, recent evidence from free air exposure systems and OTCs
2 supports results observed previously in OTC studies. For example, a series of studies
3 undertaken using free air O3 enrichment in Rhinelander, WI (Isebrands et al., 2000, 2001)
4 demonstated that O3-symptom expression was generally similar in OTCs, FACE, and also sites
5 along an ambient O3 gradient, supporting the previously observed variation among aspen clones
6 using OTCs (Karnosky et al., 1999). As has been observed in previous criteria documents, root
7 growth often is found to be the most sensitive biomass response indicator to O3.
8 Results reported since 1996 support the conclusion of the 1996 AQCD (U.S.
9 Environmental Protection Agency, 1996) that deciduous trees are generally less sensitive to
10 O3 than are most annual plants, with the exception of a few very sensitive genera such as
11 Populus and sensitive species such as black cherry. However, the data presented in Table 9-19
12 suggest that ambient O3 exposures that occur in the United States can potentially reduce the
13 growth of seedlings of deciduous species. Results from multiple-year studies sometimes show a
14 pattern of increased effects in subsequent years. In some cases, however, growth decreases due
15 to O3 may become less significant or even disappear over time. While some mature trees show
16 greater O3 sensitivity in physiological parameters such as net photosynthetic rate than do
17 seedlings, these effects may not translate into measurable reductions in biomass growth.
18 However, because even multiple-year experiments do not expose trees to O3 for more than a
19 small fraction of their life span, and because competition may in some cases exacerbate the
20 effects of O3 on individual species, determining effects on mature trees remains a significant
21 challenge.
22 In Europe, a Level I critical level has been set for forest trees based on OTC studies of
23 European beech seedlings. A critical level was defined as an AOT40 value of 10 ppm-h for
24 daylight hours for a 6-month growing season (Karenlampi and Skarby, 1996). However, other
25 studies show that some species such as silver birch may be more sensitive to O3 compared to
26 beech (Paakkonen et al., 1996).
27 For evergreen tree species, as for other tree species, the O3 sensitivity of different
28 genotypes and different species varies widely. Based on studies with seedlings in OTCs, major
29 species in the United States are generally less sensitive than are most deciduous trees, and
30 slower- growing species are less sensitive than are faster-growing species. Interacting stresses
31 such as competition stress may increase the sensitivity of trees to O3. As for deciduous species,
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1 most experiments with evergreen species have only covered a small portion of the life span of a
2 tree and have been conducted with seedlings, making estimating effects on mature trees difficult.
3 For all types of perennial vegetation, cumulative effects over more than one growing
4 season may be important, furthermore, studies for only a single season may underestimate
5 effects. Mature trees may be more or less sensitive to O3 than are seedlings, depending on the
6 species, but information on physiological traits may be used to predict some such differences.
7 In some cases, mature trees may be more sensitive to O3 than seedlings due to differences in
8 their gas exchange rates, growth rates, greater cumulative exposures, or due to the interaction of
9 O3 stress with other stresses.
10
11
12 9.7 EFFECTS OF OZONE EXPOSURE ON NATURAL ECOSYSTEMS
13 9.7.1 Introduction
14 The preceding section on species-level responses (9.6) provides a lead-in to address the
15 response of ecosystems to ozone (O3). The conclusion of the 1996 O3 AQCD was that aside
16 from the results from the San Bernardino Forest, there was no direct evidence that O3 is altering
17 natural ecosystems in the United States. This conclusion is generally valid today, except that our
18 understanding of the effects of O3 in the San Bernardino forest has been tempered by additional
19 understanding of the complicating role that N deposition plays in this system. Despite the lack
20 of any new, direct information linking O3 with ecosystem changes, numerous publications since
21 1996 have highlighted ways in which O3 may affect ecosystem structure and/or function. This
22 section addresses new and (where appropriate) older literature in order to illustrate possible
23 shifts in energy or material flow through ecosystems as a result of O3 exposure.
24 An ecosystem is defined as comprising all of the organisms in a given area interacting with
25 the physical environment, so that a flow of energy leads to a clearly defined trophic structure,
26 biotic diversity, and cycling of materials between living and nonliving parts (Odum, 1963).
27 Individuals within a species and populations of species are the building blocks from which
28 communities and ecosystems are constructed. Classes of natural ecosystems are distinguished
29 by their dominant vegetation form, e.g., tundra, wetland, deciduous forest and conifer forest.
30 Boundaries of ecosystems are delineated when an integral unit is formed by the physical and
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1 biological parts. Defined pathways for material transport and cycling, and for the flow of energy
2 are contained within the integrated unit.
3 Each level of organization within an ecosystem has functional and structural
4 characteristics. At the ecosystem level, functional characteristics include but are not limited to
5 energy flow; nutrient, hydrologic, and biogeochemical cycling; and maintenance of food chains.
6 The sum of the functions carried out by ecosystem components provides many benefits to
7 mankind as in the case of forest ecosystems (Smith, 1992). These include food, fiber production,
8 aesthetics, genetic diversity, and energy exchange.
9 Ecosystems are functionally highly integrated. Thus, changes in one part of an ecosystem,
10 such as the primary producer component, may have consequences for connected parts, such as
11 the consumer and decomposer components. For example, when needles are shed prematurely as
12 a result of O3 exposure, successional development of phyllosphere fungi inhabiting the surface of
13 ponderosa pine needles may be truncated (Bruhn, 1980). In addition, decomposer populations in
14 the litter layer may be capable of higher rates of decomposition, because the younger age classes
15 of needles falling from O3-damaged pines have higher nitrogen (N) content (Fenn and Dunn,
16 1989). Because ecological systems integrate the effects of many influences, the resulting effect
17 of O3 exposure may depend on co-occuring influences that predispose an ecosystem to stress
18 (Colls and Unsworth, 1992). One important change in our thinking since the 1996 O3 AQCD is
19 that at high levels of O3 exposure that are known to result in detectable plant responses
20 (> 250 ppm h accumulated over a growing season), N deposition must also be considered as a
21 concurrent stressor. Changes in N cycling and compartmentalization in the ecosystem result
22 from both O3 exposure and increased N deposition.
23 The vast majority of O3 effects literature addresses individual species responses (see
24 Section 9.7.4.3), which was also true in 1996 (U.S. Environmental Protection Agency, 1996).
25 This section differs from the preceding one in that physiological stress of individual species is
26 considered only within the context of the natural ecosystem. Changes in function at the level of
27 the individual are propagated through higher levels of organization, resulting in changes in
28 ecosystem structure and function. However, since ecosystem-level responses result from the
29 interaction of organisms with one another and with their physical environment, it takes longer
30 for a change to develop to a level of prominence that can be identified and measured. The
31 paucity of scientific literature on O3 effects at the ecosystem level is a result of both the
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1 complexity of ecological systems, and long response times. In addition, "indirect" effects of O3
2 on plants (e.g., affecting the plants' ability to integrate environmental stresses) may be more
3 important than direct effects on photosynthesis and respiration at the leaf level (Johnson and
4 Taylor, 1988).
5 A conceptual framework (see Table 9-22) suitable for organizing discussion of the effect of
6 O3 on ecosystems was developed by the EPA Science Advisory Board (Young and Sanzone,
7 2002). Their six essential ecological attributes (EEAs) include landscape condition, biotic
8 condition, chemical and physical characteristics, ecological processes, hydrological and
9 geomorphological processes, and natural disturbance regimes (listed as subsection headings
10 below and in Table 9-22). The major ecological effects, and gaps in our knowledge of O3
11 exposure effects at the ecosystem level are summarized at the end of this chapter. While the
12 main focus is O3 effects newly described since the last Ozone Criteria Document (EPA, 1996),
13 many key historical papers are cited to demonstrate ecosystem response, particularly where they
14 remain the only examples in the literature. Although the vast majority of published studies focus
15 on individuals, five field examples and one FACE experiment have measured several ecosystem
16 components simultaneously to better understand ecosystem response to O3. We provide an
17 overview of these six studies up-front because they provide a context for the subsequent
18 discussion on possible ecosystem effects.
19
20 9.7.2 Case Studies
21 9.7.2.1 Valley of Mexico
22 The first evidence of air pollution impacts on vegetation in the Valley of Mexico (Mexico
23 City Air Basin) were observations of foliar injury symptoms in bioindicator plants attributed to
24 O3, PAN, SO2 and possibly other pollutants (de Bauer, 1972). Subsequently, O3 injury to foliage
25 and crowns of pine trees were reported in forests to the south and southwest of Mexico City
26 (Krupa and de Bauer, 1976; de Bauer and Hernandez-Tejeda, 1986). Ozone is considered to be
27 the pollutant with the most severe impacts on vegetation within the Mexico City urban zone and
28 in forests downwind of the city. P. hartwegii is the most O3-sensitive pine species and is
29 severely impacted by high O3 exposures encountered to the south/southwest of the metropolitan
30 area (Miller et al., 2002). The potential for O3 injury is particularly high in this area, because O3
31 levels are high during the summer rainy season when soil moisture availability and stomatal
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Table 9-22. Essential Ecological Attributes for Natural Ecosystems Affected by O3
Category
Landscape Condition
• Habitat Types
Biotic Condition
• Ecosystems and
Communities
Community Extent
Community
Composition
Trophic Interactions
Insects
Diseases
Community
Dynamics
Species
Mixed conifer forest
Pinus ponderosa
Grassland communities
Coastal sage scrub
Early successional plant
community
Populus tremuloides
& Betula papyrifera
Pinus ponderosa
Pinus taeda
Pinus ponderosa
Pinus ponderosa
Populus tremuloides
Populus tremuloides
Populus hybrids
Populus hybrids
Populus tremuloides
Populus tremuloides
Picea abies &
Picea sitchensis
Pinus ponderosa
Pinus taeda
Pinus sylvestris/
mycorrhizae
Pinus ponderosa/Abies
concolor/calocedrus
decurrens
Populus tremuloides
Pinus ponderosa/
Elymus glaucus
Pinus taeda/diverse
community
Condition Measures
Community composition,
Stand structure
Relative abundance
Species composition
Species cover, richness,
equitability
Species richness, diversity,
evenness
Soil microbial community
Soil microbial community
Fungal morphotypes
Bark beetle severity
Bark beetle productivity and
predator/parasitoid density
Blotch leaf miner
performance
Aphid/natural enemy
abundance
Septoria occurrence
Rust occurrence
Rust occurrence
Forest tent caterpillar/
paratisoid performance
Needle fungi
Root disease x O3 interactions
Canker dimensions
Disease susceptibility
Abundance
Competitive status
O3 sensitivity
Tree growth
References
Miller etal. (1989);
Miller and McBride (1999)
Miller (1973);
Arbaugh et al. (2003)
Ashmore and Ainsworth (1995);
Ashmoreetal. (1995)
Westman( 1979, 1981)
Barbo etal. (1998)
Phillips et al. (2002)
Scagel and Andersen (1997)
Edwards and Kelly (1992);
Qui etal. (1993)
Cobb etal. (1968)
Dahlsten et al. (1997)
Kopper and Lindroth (2003a)
Percy et al. (2002)
Woodbury et al. (1994)
Beare etal. (1999)
Kamosky et al. (2002);
Percy et al. (2002)
Holton et al. (2003)
Maganetal. (1995)
Fenn etal. (1990)
Carey and Kelly (1994)
Bonello etal. (1993)
Minnichetal. (1995)
McDonald et al. (2002)
Andersen et al. (2001)
Barbo et al. (2002)
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Table 9-22 (cont'd). Essential Ecological Attributes for Natural Ecosystems
Affected by O3
Category
Species
Condition Measures
References
Species and
Populations
Population Size
Genetic Diversity/
Population Structure
Population
Dynamics
Pinus strobus
Firms ponderosa
Lupinus bicolor
Populus tremuloides
Trifolium repens
Plantago major
Trifolium repens
Plantago major
Mortality
Mortality
% population sensitive
% population sensitive
% population sensitive
% population sensitive
Adaptation
Population changes over time
Kamosky(1981)
Carroll et al. (2003)
Dunn (1959)
Berrang etal. (1986, 1989, 1991)
Heagleetal. (1991)
Reiling and Davison (1992a);
Davison and Reiling (1995);
Lyons etal. (1997)
Heagleetal. (1991)
Davison and Reiling (1995)
Organism Condition
• Visible Symptoms
Pinus ponderosa
Pinus Jeffreyi
Prunus serotina
Sassafras albidum
Populus nigra,
Fraxinus excelsior &
Prunus avium
Fagus sylvatica
Fraxinus americana
Grassland species
Herbaceous species
Asclepias exaltata
Rudbeckia laciniata &
Verbesina occidentalis
Asclepias incarnata
Foliar symptoms
Foliar symptoms
Foliar symptoms
Liriodendron tulipfera Foliar symptoms
Foliar symptoms
Foliar symptoms
Foliar symptoms
Foliar symptoms
Foliar symptoms
Foliar symptoms
Foliar symptoms
Foliar symptoms
Foliar symptoms
Grulke and Lee (1997);
Arbaughetal. (1998;
Salardino and Carroll (1998);
Temple etal. (1992)
Patterson and Rundel (1995);
Salardino and Carroll (1998);
Grulke et al. (2003b)
Fredericksen et al. (1995, 1996);
Chappelka et al. (1997, 1999a,b);
Hillebrandetal. (1996);
Ghosh etal. (1998);
Leeelal. (1999);
Ferdinand et al. (2000, 2003);
Yuska et al. (2003)
Somersetal. (1998);
Hillebrand etal. (1996);
Chappelka et al. (1999a)
Chappelka etal. (1999a)
Novak et al. (2003)
Gerosa et al. (2003);
Vollenweider et al. (2003a)
Schaub et al. (2003);
Ferdinand et al. (2003)
Bungener et al. (1999a)
Bergmann et al. (1999)
Chappelka etal. (1997)
Chappelka et al. (2003)
Orendovici et al. (2003)
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Table 9-22 (cont'd). Essential Ecological Attributes for Natural Ecosystems
Affected by O3
Category
Species
Condition Measures
References
Physiological Status
Reproductive Status
Pinus halepensis
Populus tremuloides
Betula pendula
Fagus sylvatica
Populus tremuloides
Pinus ponderosa
Festiva ovina
Betula pubescens
Fagus sylvatica
Populus tremuloides x
P. tremula
Prunus serotina
Fraxinus excelsior
Populus tremuloides
Pinus ponderosa
Betula pendula
Betula pendula
Acer saccharum
Apocynun
androsaemifolium
Buddleia davidii
Rubus cuneifolius
Plantago major,
Allometry
Crown architecture
Crown architecture
Crown architecture
Root dry weight
Root/shoot ratio
Root/shoot ratio
Root/shoot ratio
Root/shoot ratio
Root/shoot ratio
Leaf area
Leaf area
Leaf area index
Carbon allocation to
mycorrhizae
Decreased winter bud
formation
Delayed bud break
Early bud break
Flowering time
Flowering time
Flowering time
Pollen germination &
Pollen tube elongation
Fragaria x ananassa Fruit yield
Plantago major Seed yield
Understorey herbs Seed yield
Welburn and Wellburn (1994)
Dicksonetal. (2001)
Kull et al. (2003)
Stribley and Ashmore (2002)
Colemanetal. (1996)
Grulkeetal. (1998);
Grulke and Balduman, (1999)
Warwick and Taylor, (1995)
Mortensen(1998)
Paludan-Muller et al. (1999);
Landollelal. (2000)
Oksanenetal. (200la)
Neufeld et al. (1995)
Wiltshire et al. (1996)
Karnosky et al. (2003a)
Andersen and Rygiewicz (1995)
Oksanen (2003a,b)
Prozherina et al. (2003)
Bertrand et al. (1999)
Bergweiler and Manning,
(1999)
Findleyetal. (1997)
Chappelka (2002)
Stewart (1998)
Drogoudi and Ashmore
(2000,2001)
Reiling and Davison (1992b);
Pearson et al. (1996); Whitfield
et al. (1997); Lyons and Barnes
(1998)
Harward and Treshow (1975)
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Table 9-22 (cont'd). Essential Ecological Attributes for Natural Ecosystems
Affected by O3
Category
Species
Condition Measures
References
Ecological Processes
• Energy Flow
Primary Production Pinus ponderosa
Pinus ponderosa
Populus tremuloides
Betula pendula
Betula pendula
Quercus rubra
Populus tremuloides
Populus tremuloides
Pinus ponderosa
Quercus rubra
Populus tremuloides
Pinus ponderosa
Betula pendula
Betula pendula
Fragaria vesca
Pinus taeda
Lespedeza cuneata &
Schizacbyrium
scoparium
Liriodendron tulip/era
Prunus serotina
Pinus jeffreyi
Pinus ponderosa
Pinus strobus
Pinus taeda
Photosynthesis
Needle retention
Photosynthesis
Photo sy nthe sis/conductance
Stem respiration & radial
growth
Root turnover
Soil respiration
Soil respiration
Soil respiration
Carbon partitioning &
allocation
Carbon allocation
Carbon allocation
Carbon allocation
Carbon allocation
Carbon allocation
Root respiration
Yield
Radial growth
Radial growth
Radial growth
Radial growth (no effect)
Radial growth
Radial growth
Miller etal. (1969);
Clark etal. (1995);
Takemoto et al. (1997);
Grulke et al. (2002b)
Temple etal. (1993)
Colemanetal. (1995b);
Noormets et al. (2001a,b);
Sharmaetal. (2003;
Karnosky et al. (2003a)
Oksanen (2003a,b)
Matyssek et al. (2002)
Keltingetal. (1995)
Colemanetal. (1996)
King etal. (2001)
Andersen and Scagel (1997);
Scagel and Andersen (1997),
Andersen (2000)
Samuelson and Kelly (1996)
Colemanetal. (1995b)
Andersen et al. (1997);
Grulke etal. (1998);
Grulke and Balduman (1999);
Grulke etal. (2001)
Karlsson el al. (2003)
Oksanen and Saleem, (1999);
Salem etal. (2001)
Manninen et al. (2003)
Edwards (1991)
Powell et al. (2003)
Somers etal. (1998)
Vollenweider et al. (2003b)
Peterson etal. (1987)
Peterson etal. (1993)
Bartholomay et al. (1997)
and Downing (1995, 1996)
Braunetal. (1999)
Fagus sylvatica
Picea abies
Populus tremuloides
Pinus ponderosa
Stem volume
Stem volume
Volume growth
Root growth
Wallm et al. (2002)
Isebrands et al. (2001)
Andersen et al. (1991)
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Table 9-22 (cont'd). Essential Ecological Attributes for Natural Ecosystems
Affected by O3
Category
Species
Condition Measures
References
Net Ecosystem
Production
Material Flow
Organic Carbon
Cycling
Nitrogen Cycling
Other Nutrient
Cycling
Northern hardwoods
Northern hardwoods
Growth Efficiency Plantago major
Grassland species
Native herbs
Grasses and herbs
Populus tremuloides
Fagus sylvatica
Picea abies
Primus serotina
Populus tremuloides &
Be tula papyri/era
Andropogon virginicus
& Rubus cuneifolius
Liriodendron tulipera
Populus deltoides
Pinus ponderosa
Pinus sylvestris
Pinus ponderosa
Pinus taeda
Prunus serotina &
Liriodendron tulip/era
Picea sitchensis &
Pinus sylvestris
Pinus ponderosa
Hydrology and Geomorphology
• Water Budget Picea rubens
Pinus armandi
Pinus jeffreyi
Picea abies
Fraxinus excelsior
Be tula pendula
Populus hybrids
NPP estimates
Biomass estimates
Relative growth rate
Relative growth rate
Relative growth rate
Relative growth rate
Relative growth rate
Relative growth rate
Relative growth rate
Relative growth rate
Altered foliar C:N ratio and
N resorption efficiency
Litter decomposition rate
Litter decomposition rate
Litter decomposition rate
Litter decomposition rate
Litter decomposition
(no effect)
Altered foliar N
Foliar N & O3 exposure
(no effects)
Altered foliar N metabolism
Altered foliar N
Foliar leaching (no effect)
Nutrient availability & O3
Water-use efficiency
(no effect)
Water-use efficiency
Canopy transpiration
Transpiration, xylem sap flow
Stem flow of water
Water-use efficiency
Water-use efficiency
Laurence et al. (2000)
Hogsettetal. (1997)
Reiling and Davison (1992a);
Davison and Reiling (1995);
Lyons et al. (1997);
Davison and Barnes (1998)
Bungener et al. (1999b)
Warwick and Taylor (1995)
Pleijel and Danielsson (1997)
Yun and Laurence (1999)
Bortier et al. (2000)
Karlsson et al. (2002)
Lee et al. (2002)
Lindroth et al. (2001)
Kim etal. (1998)
Scherzer and Rebbeck (1998)
Findlay and Jones (1990)
Fenn and Dunn (1989)
Kainulainen et al. (2003)
Momen and Helms (1996)
Bytnerowicz et al. (1990)
Mandersheid et al. (1992)
Boerner and Rebbeck (1995)
Skeffington and Sutherland
(1995)
Bytnerowicz et al. (1990)
Laurence et al. (1997)
Shan etal. (1996)
Grulke et al. (2003a)
Maier-Maercker (1997)
Wiltshire et al. (1994)
Maurer and Matysek (1997)
Reich and Lassoie (1984)
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Table 9-22 (cont'd). Essential Ecological Attributes for Natural Ecosystems
Affected by O3
Category
Species
Condition Measures
References
Natural Disturbance Regimes
• Frequency Pinus ponderosa
Intensity
Extent
Duration
Pinus ponderosa
Picea sitchensis
Pinus halepensis
Picea rubens
Fagus sylvatica
Picea abies
Pinus ponderosa
Pinus ponderosa
Pinus ponderosa
Frequency of fire
Occurrence of bark beetle
outbreaks
Winter damage
Reduced winter damage
Freezing tolerance
Drought stress
Drought stress
Fire intensity
Extent of bark beetle attack
Duration of bark beetle attack
McBride and Laven (1976);
Minnichetal. (1995);
Miller and McBride (1999)
Pronosetal. (1999);
Dahlsten et al. (1997)
Lucas etal. (1988)
Wellburn and Wellburn, (1994)
Waiteetal. (1994)
Pearson and Mansfield
(1993, 1994)
Maier-Maercker (1998);
Maier-Maercker and
Koch (1992)
Miller and McBride (1999)
Minnichetal. (1995)
Minnichetal. (1995)
Source: Young and Sanzone (2002.
1 conductance are greatest and these factors enhance O3 uptake and injury. Decline of A. religiosa
2 (oyamel) in the Desierto de los Leones National Park is a well-known example of dramatic
3 dieback and mortality of entire forest stands due primarily to air pollution stress
4 (Alvarado-Rosales and Hernandez-Tejeda, 2001). Other factors, such as a lack of stand
5 thinning, also contribute to forest decline. Lead in automobile gasoline was phased out in 1990,
6 and foliar concentrations of heavy metals in forest species are not now at phytotoxic levels (Fenn
7 et al. 2001a). Sulfur dioxide concentrations decreased in the early 1990s as a result of regulatory
8 mandates limiting their emissions. Sensitive plants in the NE and NW sectors of the Mexico
9 City urban zone where concentrations are highest may still be impacted by exposure to ambient
10 SO2 levels. Deposition of ionic forms of N and S are high in forested areas southwest of Mexico
11 City. The effects of these chronic nutrient inputs to the forest are only beginning to be
12 investigated and understood.
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1 The ecological perturbations caused by severe air pollution exposures in forests located
2 downwind of Mexico City are expected to continue for the near future (the next 5 to 10 years),
3 largely as a result of high O3 concentrations, as well as of N oxides emissions. The longer-term
4 response is more uncertain, and depends largely on the effectiveness of regulatory emissions
5 control strategies. Currently, pollutant levels are declining. Forest responses to this trend will
6 depend on how long it takes to reduce levels sufficiently to allow sensitive species to recover.
7 Some of the change to the ecosystem is probably irreversible, such as the loss of lichen diversity
8 and of other O3-sensitive species (Zambrano et al., 2002).
9
10 9.7.2.2 San Bernardino Mountains
11 The San Bernardino Mountains lie east of the Los Angeles Air Basin (California, USA),
12 and significant levels of pollution have been transported into the mountain range, including a
13 Class I Wilderness area. The effects of O3 exposure on the mixed conifer forest of the
14 San Bernardino Mountains is perhaps the longest and most thoroughly documented O3
15 ecological effects evaluation (Miller and McBride, 1999). In this classic case study linking
16 tropospheric O3 exposure to damage to an entire forest ecosystem (U.S. Environmental
17 Protection Agency, 1996) (Table 9-23), Miller et al. (1963) first identified the unique foliar
18 chlorotic mottle that was occurring on two of the dominant tree species, Pinus ponderosa and
19 P.jeffreyi. Levels of O3 averaging 100-120 ppb over 24 hours with 1-hour peaks well into the
20 200 ppb range were common in the region in the 1960s and 1970s (Miller and McBride, 1999).
21 Single hour peak values have declined in recent years due to heavily regulated pollution control,
22 but O3 concentrations in the moderate range continue to rise and accumulate, increasing the
23 overall cumulative exposure (Arbaugh et al., 1998; Takemoto et al., 2001; Lee et al., 2003).
24 Since the 1996 O3 AQCD (U.S. Environmental Protection Agency, 1996) was written, the
25 concurrent role of N deposition in modifying ecosystem response to O3 exposure in the San
26 Bernardino Mountains has been further elucidated (Fenn et al., 1996, 2003; Bytnerowicz et al.,
27 1999; Takemoto et al., 2001; Bytnerowicz, 2002). Both O3 exposure and N deposition reduce
28 foliar retention (Grulke and Balduman, 1999) and alter tissue chemistry of both needles and litter
29 (Poth and Fenn, 1998). In addition, confounding factors such as drought and fire suppression
30 add to the complexity of ecosystem response (Minnich et al., 1995; Takemoto et al., 2001;
31 Arbaugh et al., 2003). Extensive crown injury measurements have also been made, linking
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Table 9-23. Case Studies Demonstrating the Ecological Effects of O3
Study
Keystone
Species
Period
Study Type Studied Key Ecological Findings
Valley of
Mexico
Pinus hartwegii,
Abies religiosa
Field transects
San Bernardino Pinus ponderosa, Field transects
Mountains P. jeffreyi
35 yrs • Significant foliar injury
(de Bauer, 1972; Krupa and Bauer, 1976;
Hernandez-Tejeda, 1986)
• Community composition changes
(Alvarado-Rosales and Hernandez-
Tejeda, 1986)
• Species richness changes
(Zambrano et al., 2002)
40 yrs • Community composition changes
(Miller, 1973; Minnichet al., 1995;
Arbaughetal.,2003)
• Population changes
(McBride and Laven, 1999)
• O3/pine^ark beetle interaction
(Pronosetal. 1999)
• Altered C flows
(Grulke et al., 2002b; Grulke et al., 1998;
Grulke andBalduman, 1999; Grulke
et al., 2001; Arbaugh et al., 1999)
• Interaction of O3, drought, N deposition:
(Fennetal., 1996; Grulke, 1999;
Takemoto et al., 2001)
• Altered carbon cycling
(Arbaugh et al., 1999)
Sierra Nevada
Mountains
Pinus ponderosa,
P. jeffreyi
Appalachian
Mountains
Fraxinus
americana,
Liriodendron
tulip/era,
Pinus strobus,
Prunus serotina
Field 35 yrs • Wide-scale nature of effects (Miller and
Millecan, 1971; Miller etal,. 1972)
• Link to decreased growth
(Peterson et al., 1987, 1991, 1995)
• Quantification of O3 flux
(Bauer et al., 2000; Panek et al., 2002;
Goldstein etal., 2003)
• Cumulative O3 effects
(Takemoto et al., 1997)
• Canopy level responses
(Grulke etal., 2003a,b)
• Population changes (Carroll et al., 2003)
Field 25 yrs • Link of visible symptoms to growth
decreases (McLaughlin et al., 1982;
Somersetal., 1998)
• Wide-scale nature of effects (Chappelka
et al., 1999a; Hildebrand et al., 1996)
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Table 9-23 (cont'd). Case Studies Demonstrating the Ecological Effects of O3
Study
Aspen FACE
Keystone
Species
Acer sacchamm,
Betula
papyri/era,
Populus
tremuloides
Period
Study Type Studied Key Ecological Findings
Open-air O3 6 yrs • Competitive interactions
exposure (McDonald et al., 2002)
• O3/aspen/rust interaction
(Karnoskyetal, 2002)
• Plant- insect interactions (Percy et al.,
2002; Holtonetal, 2003)
• C & N cycling (Lindroth et al., 2001;
King etal., 2001)
• Moderation of CO2 responses by O3
(Isebrands et al., 2001; Wustman et al.,
2001; McDonald et al., 2002;
Karnoskyetal.,2003a)
Plantago Plantago major Field 20 yrs • Population structure (Davison and
Reiling, 1995; Lyons et al., 1997)
• O3 resistance
(Reiling and Davison, 1992a)
• Adaptation (Davison and Reiling, 1995)
Carpathian Pinus sylvestris, Field 15 yrs • Significant foliar injury
Mountains Picea abies • Community composition changes
• Species diversity changes
1 ambient O3 exposure data to chlorotic mottle and fascicle retention (Arbaugh et al., 1998).
2 Ozone exposure and N deposition reduce carbon allocation to stems and roots (Grulke et al.,
3 1998; 2001), further predisposing trees to drought stress, windthrow, root diseases, and insect
4 infestation (Takemoto et al., 2001). Increased mortality of susceptible tree species (ponderosa
5 and Jeffrey pine, Pinus ponderosa, P. Jeffreyii) has shifted community composition towards
6 white fir and incense cedar (Abies concolor, Libocedrus decurrens) and has altered forest stand
7 structure (Miller et al., 1989). Ozone exposure is implicated in projected changes in stand
8 composition (McBride and Laven, 1999) toward a predominance of oaks, rather than mixed
9 conifer forests. Forest understory species have also been affected (Temple, 1999). These
10 individual species responses collectively have affected trophic structure and food web dynamics
11 (Dahlsten et al., 1997; Pronos et al., 1999), as well as C and N cycling (Arbaugh et al., 2003)
12 (Table 9-24). Because of the high N deposition in the San Bernardino Mountains, it is difficult
13 to separate out effects of only O3 versus those due to the combined effects of O3 and N
14 deposition (Table 9-25).
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Table 9-24. The Most Comprehensively Studied Effects of O3 on Natural Ecosystem is
the San Bernardino Mountain Forest Ecosystem. Citations Focus on Research
Published Since U.S. EPA (1996).
Pollutant Occurrence
Reference
O3 exposures N deposition
Cellular, Biochemical
Foliar pigments
Antioxidants
Foliar Responses
Foliar Symptoms
Gas Exchange
Photosynthesis & Conductance
O3flux
Foliar nutrients
Whole Organism
Growth/Biomass
• Aboveground
• Belowground
• Root/shoot ratio
• Carbon allocation
• Crown vigor
Ecosystem
Community dynamics/succession
Simulations
Understory vegetation
Pest interactions
• Bark beetle/predators
• Disease occurrence
• Litter decomposition
Disturbance
• Bark beetle occurrence
• Fire frequency
Grulke et al. (1998); Fenn et al. (1996, 2000,
2003); Bytnerowicz et al. (1999)
Grulke and Lee (1997); Grulke (1999); Tausz
et al. (1999a,b 2001); Tausz et al. (1999a,b,c,
2001)
Grulke and Lee (1997); Arbaugh et al. (1998);
Miller and Rechel (1999)
Grulke (1999); Grulke and Retzlaff (2001);
Grulke et al. (2002a,b)
Grulke and Balduman (1999)
Grulke et al. (1998); Grulke and Balduman
(1999)
Grulke and Balduman (1999)
Grulke etal. (2001)
Arbaugh et al. (1998); Miller and Rechel (1999)
McBride and Miller (1999); Arbaugh et al. (2002)
Arbaugh etal. (1999)
Temple (1999)
Dahlsten et al. (1997); Pronos et al. (1999)
Miller and Rechel (1999); Pronos et al. (1999)
Minnich et al. (1995); Minnich (1999); Miller and
McBride (1999)
January 2005
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Table 9-25. Effects of Ozone, Ozone and N Deposition, and Ozone and Drought Stress
on Pinus ponderosa and Pinus jeffreyi in the Sierra Nevada and the San Bernardino
Mountains, California. Citations are Focused on Research Published
Since U.S. EPA (1996).
Foliar Biochemistry and
Tissue Chemistry
Total ascorbate
Dehydroascorbate
Total glutathione
Oxidized glutathione
a Carotenoids
Foliar nitrogen
C:N ratio of foliage2
Starch
Chlorophyll content
Gas Exchange
A,,,^ lower canopy
Amax whole canopy
A^ seedlings
Stomatal limitation
Stomatal conductance
Foliar respiration
O3flux
O3 O3 + N deposition
d1 d
i n.d.
d i
i i
i n.d.
d i
i n.d.
ad. d
d id
n.d. i
d ad.
di n.d.
a to ps n.d.
d di
as. i
d as.
O3 + Drought
i
d
d
d
d
d
d
i
d
d
d
ad.
i
d
d
d
References
Tausz et al. (2001); Grulke et al.
(2003b)
Grulke et al. (2003b)
Tausz etal. (2001)
Tausz etal. (2001)
Grulke et al. (2003b)
Grulke and Lee (1997);
Grulke etal. (1998);
Poth and Fenn (1998)
Poth and Fenn (1998);
Grulke et al. (2003b)
Grulke etal. (2001)
Takemoto et al. (1997); Grulke
and Lee (1997; Grulke et al.
(1998, 2003b); Grulke (1999);
Tausz etal. (2001)
Grulke (1999); Grulke and
Retzlaff (2001); Grulke et al.
(2002b); Panek (2004)
Panek and Goldstein (2001);
Grulke et al. (2003b); Misson
et al. (2004)
Grulke and Retzlaff (2001)
Panek and Goldstein (2001)
Grulke (1999; Grulke et al.
(2003a); Panek (2004)
Grulke et al. (2002a; Grulke
(1999)
Panek and Goldstein (2001);
Panek et al. (2002; 2003); Grulke
et al. (2002b; 2004)
January 2005
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Table 9-25 (cont'd). Effects of Ozone, Ozone and N Deposition, and Ozone and
Drought Stress on Pinus ponderosa and Pinus jeffreyi in the Sierra Nevada and the
San Bernardino Mountains, California. Citations are Focused on Research
Published Since U.S. EPA (1996).
Growth and Productivity
Foliar biomass
Height growth
Bole diameter growth
Fine root biomass
03
ad.
ad.
d
d
O3 + N deposition
i
i
i
d
O3 + Drought
d
d
d
i
References
Grulke and Balduman (1999)
Grulke and Balduman (1999)
Grulke and Balduman (1999)
Grulke etal. (1998)
Leaf Surfaces
Stomatalocclusion i n.d. n.d. Bytnerowiczetal. (1999);
Bytnerowicz and Turunen (1994)
Trophic Interactions
Bark beetle n.s. i i Pronos etal. (1999)
Ecosystem Level
Competitive indices n.d. d i Miller and Rechel (1999)
'Responses are shown as significant increases (i), significant decreases (d), both significant decreases and
increases reported (di), nonsignificant effects (n.s.), and no data (n.d.) compared to trees or seedlings at field sites
with lower ozone, drought stress, or lack of significant N deposition (< 10 kg ha"1 yf'). Frequently n.d. was used
for lack of a control site without compounding high N deposition. Foliar analyses and leaf surface properties
were largely determined from previous year needles. Gas exchange data were generally from previous year
needles at peak growing season, prior to late summer drought (mid to late July).
Abbreviations: C = carbon; N = nitrogen; A^ = maximum photosynthesis rate.
1 9.7.2.3 Sierra Nevada Mountains
2 The western slope of the Sierra Nevada Mountains in central and southern California has
3 also been exposed to elevated O3 for a long time, although the effects have been much less than
4 those observed in the San Bernardino Mountains. Symptoms of O3 injury have been found on
5 ponderosa and Jeffrey pines in all of the Sierra Nevada National forests and parks (Carroll et al.,
6 2003). First identified as a problem in the 1970s (Miller et al., 1972), elevated O3 with daytime
7 means of 60-80 ppb are common (Bohm et al., 1995; Bauer et al., 2000; Bytnerowicz et al.,
8 2002a; Panek et al., 2002). The west-slope Sierra Nevada forests are also exposed to a wide
9 range of additional gaseous and paniculate pollutants, including various S and N compounds
January 2005 9-268 DRAFT-DO NOT QUOTE OR CITE
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1 (Bytnerowicz et al., 1999; Takemoto et al., 2001; Fenn et al., 2003), but at levels much lower
2 than in the San Bernardino Mountains. Typical O3-induced visible foliar symptoms, including
3 chlorotic mottle, chlorophyll degradation, and premature senescence, are commonly found on
4 sensitive genotypes ofponderosapine (Peterson et al., 1991; Arbaugh et al., 1998; Staszak et al.,
5 2003) and Jeffery pine (Peterson et al., 1987; Patterson and Rundel, 1995; Arbaugh et al., 1998;
6 Grulke et al., 2003b). Other important conifers in the region, such as giant sequoia, appear to be
7 relatively tolerant to O3 (Grulke et al., 1996). The symptoms of foliar injury and growth
8 reductions have been verified on seedlings in O3 exposure chambers (Temple, 1988; Momen and
9 Helms, 1996; Momen et al., 2002).
10 Ozone foliar injury of dominant pine species in the Sierra Nevada Mountains is correlated
11 to decreased radial growth in both ponderosa pine (Peterson et al., 1991) and Jeffreyipine
12 (Peterson et al., 1987; Patterson and Rundel, 1995). Because of the large amount of intraspecific
13 variation in O3 sensitivity in these two species, O3 exposure may be a selective agent (Patterson
14 and Rundel, 1995), with differential mortality rates for sensitive individuals (Carroll et al.,
15 2003). The region's forests may also be experiencing subtle changes in species composition and
16 community structure (Patterson and Rundel, 1995; Takemoto et al., 2001).
17 Based on fire scar dating, reconstructions of stand age classes, historical records, and
18 present stand structure, fire has been largely excluded in western forests for the last 75 to 100
19 years (Minnich, 1999; Minnich and Padgett, 2003). Fire exclusion has resulted in fewer, large,
20 stand-replacing fires rather than a mosaic of smaller, lower intensity fires. The change in fire
21 intensity may have selectively altered stand structure, fitness and competitiveness of component
22 species, and their susceptibility to atmospheric pollutants and other stressors (Minnich, 1999).
23 Short-lived (50-80 years) species such as knobcone and Coulter pine, which occur at the
24 interface between the chaparral and the mixed conifer forest, may already have been selected for
25 O3 tolerance due to seedling establishment (the most sensitive tree age class in conifers) after
26 large fires in the 1950's (Minnich, 1999). Strong measures to suppress fires have largely kept
27 chaparral fires from invading the mixed conifer forests, and stand densification in the mixed
28 conifer zone has increased. High stand density, in turn, may weaken the younger cohorts and
29 increase sensitivity to atmospheric pollution (Minnich, 1999).
30 Other disturbances that play a potential role in sensitivity to atmospheric pollution include
31 cycles of drought stress. Nearly every decade is marked by one or more years of very low
January 2005 9-269 DRAFT-DO NOT QUOTE OR CITE
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1 precipitation (Graumlich, 1993). During extended periods of drought, foliar injury is lower than
2 in subsequent years with higher average precipitation (Carroll et al., 2003). In the first several
3 years (1975-1977) of a Sierran-wide assessment of O3 injury to pines, O3 injury increased,
4 because of greater water availability due to greater stomatal conductance and, presumably,
5 greater O3 uptake. Trees instrumented with monitors to directly measure canopy transpiration
6 had 20% greater stomatal conductance in mesic microsites (riparian areas, mid-slope seeps) than
7 trees in xeric microsites (rock outcrops) (Grulke et al., 2003a). Although the Sierra Nevada
8 experienced a prolong drought between 1987 and 1993, it was less severe and O3 injury did not
9 significantly decrease (Carroll et al., 2003). The same plots showed only a slight increase in O3
10 injury between 1993 and 2000. Drought stress, in general, can make trees more susceptible to
11 insect and pathogen infestation. However, serious outbreaks of insect infestation are believed to
12 be indicators, not a cause, of existing stress in the forest (Wickman, 1992).
13
14 9.7.2.4 Appalachian Mountains
15 The southern Appalachian Mountain region experiences some of the highest O3 exposures
16 of any natural areas in the eastern United States (Mueller, 1994; Hildebrand et al., 1996;
17 Samuelson and Kelly, 1997; Chappelka et al., 1997). Since the region is the home of the
18 Shenandoah and Great Smokey Mountains National Parks, which have Class I air quality
19 designations by the 1977 Clean Air Act, there has been considerable study of the region's
20 dominant forest species to determine O3 effects. Visible foliar symptoms of O3 have been found
21 in natural ecosystems consisting of Sassafras albidum (Chappelka et al., 1999a), Prunus serotina
22 (Hildebrand et al., 1996; Chappelka et al., 1997; 1999b; Samuelson and Kelly, 1997),
23 Liriodendron tulipifera and Fraxinus Americana (Hildebrand et al., 1996; Chappelka et al.,
24 1999a). Visible foliar symptoms induced by O3 have been recreated on the same species in
25 chamber studies (Duchelle et al., 1982; Chappelka et al., 1985; Chappelka and Chevone, 1986;
26 Samuelson, 1994; Fredericksen et al., 1995). No response to O3 exposure has been found for
27 other hardwood trees, nor for the three conifer species tested (Neufeld et al., 2000).
28 Long term foliar injury symptoms have been correlated with decreased radial growth in
29 tulip poplar and cherry {Liriodendron tulipifera and Prunus serotina, Somers et al. (1998) and
30 with decreased biomass in cherry (Neufeld et al., 1995). Although climatic conditions (drought)
31 largely explained radial growth reductions, O3 exposure may have also contributed (McLaughlin
January 2005 9-270 DRAFT-DO NOT QUOTE OR CITE
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1 and Downing, 1996). Ozone exposure may also be affecting the understory vegetation in the
2 region (Duchelle and Skelly, 1981; Duchelle et al., 1983; Chappelka et al., 1997; Chappelka
3 et al., 2003; Davison et al., 2003) and community composition (Barbo et al., 1998), through
4 impacts both on growth and reproduction (Chappelka, 2002). Foliar litter from trees exposed to
5 elevated O3 have lower decomposition rates (Kim et al. 1998). Other air pollutants are likely to
6 be found in this ecosystem but not at such high deposition values found in the California studies.
7 A decline in forest health in the northern Appalachians has been primarily attributed to the
8 effects of acidic fog and rain on soil acidification, lower Ca2+ availability, reduction in fine root
9 biomass, and modification of cuticular wax. However, fog- and O3-exposed red spruce forests
10 also show winter injury (Percy, 2002).
11
12 9.7.2.5 Plantago Studies in the United Kingdom
13 One of the most well-documented studies of population and community response to O3
14 effects are the long-term studies of Plantago major in native plant communities in the United
15 Kingdom (Reiling and Davison, 1992a; Davison and Reiling, 1995; Lyons et al., 1997).
16 Sensitive populations of P. major had significant growth decreases in elevated O3 (Reiling and
17 Davison, 1992b and 1992c; Pearson et al., 1996; Whitfield et al., 1997) and reduced fitness as
18 determined by decreased reproductive success (Reiling and Davison, 1992b; Pearson et al.,
19 1996). While spatial comparisons of population responses to O3 are complicated by other
20 environmental factors, rapid changes in O3 resistance were imposed by ambient levels and
21 variations in O3 exposure (Davison and Reiling, 1995). Molecular patterns of genetic variation
22 suggest that change in O3 resistance over time probably resulted from selection on genotypes
23 already present in local populations, rather than through an influx of new P. major germplasm
24 (Wolff et al., 2000). The highest correlations between O3 resistance and ambient O3
25 concentrations occurred at the site of seed collection (Lyons et al., 1997), rather than between O3
26 resistance and other climatic variables, as was found for aspen (Berrang et al., 1991).
27
28 9.7.2.6 Forest Health in the Carpathian Mountains
29 The Carpathian Mountains cross five countries (the Czech Republic, the Slovak Republic,
30 Poland, Romania, and the Ukraine) and contain many national parks and several biosphere
31 reserves. The forests were largely cleared in the 15th century, and were then planted in Norway
January 2005 9-271 DRAFT-DO NOT QUOTE OR CITE
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1 spruce plantations. As elevation increases, beech (Fagus sylvaticd) or beech-fir (Abies alba)
2 forests grade into Norway spruce (Picea abies) or spruce-fir forests. Near the treeline, Norway
3 spruce combines with dwarf mountain pine (Pinus mugo). Dwarf mountain pine forms an
4 almost pure stand just below alpine vegetation.
5 The forests of the Carpathian Mountains have been subjected to anthropogenic stressors
6 (shepherding, metal mining, wood harvest for structures and paper) for hundreds of years, as
7 described for the Tatra Mountains in the southern Carpathians (Wezyk and Guzik, 2004). The
8 Carpathians have been subjected to regional air pollution stresses since industrialization. Most
9 of the effects of air pollution on forest health degradation were due to (1) heavy metal
10 deposition, (2) soil acidification by acid deposition, and (3) subsequent pest outbreaks, the
11 combination of which led to the forest decline and dieback between 1970 and 1989 (Dunajski,
12 2004). By the 1980's, industrial pollutants such as SO2 and heavy metals significantly declined,
13 but O3 exposure has continued to increase (Bytnerowicz et al., 2004). Increased ownership and
14 use of private cars in Central Europe, as well as long-range transport of O3 from western Europe,
15 are believed to be responsible for the continued increase in photooxidants. In 1995, drought
16 resulted in significant mortality, as well as an epidemic of bark beetle infestation in subsequent
17 years. A network of air quality monitoring sites was installed across Europe in the late 1980's, as
18 part of the International Cooperative Programme on Assessment and Monitoring of Air Pollutant
19 Effects on Forests (ICP Forests). Mean defoliation rates for six important forest species across
20 Europe have increased or remained unchanged from 1989 to 1999 (Percy, 2002). Ozone
21 concentrations experienced in the Tatra Mountains, especially along the southern slopes,
22 occasionally reach 190-200 ppb as two-week-long averages, with the highest values experienced
23 in early summer at elevations of 1700 to 2300 m (Bytnerowicz et al., 2004). In other parts of the
24 Carpathian Mountains, peak two-week averages of O3 concentrations were lower, at 160 ppb
25 (Bytnerowicz et al., 2002). For all trees inventoried, about 13% exhibited greater than 25%
26 defoliation during 1997 to 2000. There was no difference between extent of damage for
27 broadleaves or conifers. Trees in Poland and the Czech Republic are the most affected by air
28 pollution, and the least damaged forests are in Romania (Badea et al., 2002).
29 The extent to which O3 exposure affects forest health degradation, and slows forest
30 degradation, is still unknown in Europe. In many of the published studies, response to a known
31 O3 gradient is largely confounded by other pollutants and/or climatic gradients (Bytnerowicz
January 2005 9-272 DRAFT-DO NOT QUOTE OR CITE
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1 et al., 1996; Szaro et al., 2002; Widadki et al., 2004). Current levels of ambient O3 are believed
2 to be high enough to reduce bole radial growth (Percy, 2004). Although average O3
3 concentration alone was not related to bole growth, the peak hourly O3 concentration was
4 negatively correlated to growth (Muzika et al., 2002). Recent evidence indicates that canopy
5 health of European white oak (Quercus robur\ Norway spruce, maritime pine (Pinus pinaster).,
6 and beech has significantly declined (Huttunen et al., 2002). However, the canopy health of
7 Scots pine (Pinus sylvestris) has improved. The network of air quality monitoring stations and
8 forest plots is extensive and active. Subsequent correlative analyses including both
9 meteorological and air quality attributes throughout the EU will help to determine the specific
10 role of O3 exposure in forest decline. Historical effects of anthropogenic disturbance may still
11 be confounding.
12
13 9.7.2.7 Field Exposure System (FACE), Rhinelander, Wisconsin
14 The Aspen Free-Air CO2 Enrichment (FACE) facility was designed to examine the effects
15 of both elevated CO2 and O3 on aspen (Populus tremuloides\ birch (Betula papyfera), and sugar
16 maple (Acer saccharum) in a simple reconstructed plantation characteristic of Great Lakes
17 Aspen-dominated forests (Karnosky et al., 1999, 2003a). Instead of using chambers to expose
18 the plants to desired gas concentrations, the gas is piped up vertical delivery tubes in the open
19 air. The vertical delivery pipes surround a 30-m diameter circular plot with five different aspen
20 clones in half of the plot, one quarter of the plot planted in aspen and birch, and one quarter in
21 aspen and maple. The O3 treatment for the first five years was 1.5* ambient, with ambient O3
22 exposures averaging 35 to 37 ppb (12 h daytime over the growing season) compared to elevated
23 O3 rings averaging 49 to 55 ppb for the same time period (Karnosky et al., 2003a).
24 Elevated CO2, elevated O3, and elevated CO2 + O3 have had effects on most system
25 components being measured in the study (Table 9-26; Karnosky et al., 2003a). One interesting
26 finding of the project has been the nearly complete offset by elevated O3 of the enhancements
27 induced by elevated atmospheric CO2 for the pioneer keystone species P. tremuloides (Isebrands
28 et al., 2001) and B. papyri/era (Percy et al., 2002), even though O3 exposure alone did not
29 always result in a significant response when compared to controls. They also found evidence
30 that the effects on above- and below-ground growth and physiological processes have cascaded
31 through the ecosystem, even affecting microbial communities (Larson et al., 2002; Phillips et al.,
January 2005 9-273 DRAFT-DO NOT QUOTE OR CITE
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Table 9-26. Summary of Responses of Populus tremuloides to Elevated CO2
(+200 fimol mol"*), O3 (1.5 x ambient), or CO2+O3 Compared with Control During
Three Years of Treatments at the Aspen FACE Project
(Modified from Karnosky et al. 2003a)
Foliar Gene Expression
and Biochemistry
Rubisco; RbcS2 transcripts
PAL transcipts
Ace oxidase, catalase
Ascorbate peroxidase
Glutathione reductase
Phenolic glycosides
Tannins
Foliar nitrogen
C:N ratio of foliage
Starch
Gas Exchange
Amax lower canopy
A,^ whole canopy
Stomatal limitation
Stomatal conductance
Foliar respiration
Soil respiration
Microbial respiration
Stomatal density
Chlorophyll content
Chloroplast structure
O3flux
Growth and Productivity
Leaf thickness
Leaf size
Leaf area
CO2
d1
d
d
d
d
i
n.s.
d
i
d
n.s.
ii
d
d
n.s.
ii
ii
n.s.
d
i
d
i
i
i
03
d
i
i
n.s.
i
d
i
n.s.
n.s.
d
dd
dd
n.s.
di
i
d
n.s.
n.s.
d
d
ii
n.s.
d
d
CO2 + O3
dd
d
d
d
d
n.s.
i
d
ii
n.s.
id
n.s.
d
d
n.s.
n.s.
n.s.
n.s.
d
d
i
n.s.
d
n.s.
Reference
Wustman et al. (2001); Noormets et al. (200 la)
Wustmanetal. (2001)
Wustman etal. (2001)
Wustmanetal. (2001)
Wustmanetal. (2001)
Lindroth et al. (2001); Kopper and Lindroth
(2003a,b)
Lindroth et al. (2001); Kopper and Lindroth
(2003a,b)
Lindroth et al. (2001); Kopper and Lindroth
(2003a,b)
Lindroth etal. (2001)
Wustmanetal. (2001)
Takeuchi et al. (2001); Noormets et al. (200 la)
Noormets et al. (200 Ib); Sharma et al. (2003)
Noormets et al. (200 la)
Noormets et al. (200 la)
Takeuchi et al. (2001); Noormets et al. (200 Ib)
King etal. (2001)
Phillips et al. (2002)
Percy et al. (2002)
Wustmanetal. (2001)
Oksanen et al. (200 Ib); Takeuchi et al. (2001);
Wustmanetal. (2001)
Noormets et al. (200 la)
Oksanen etal. (200 Ib)
Wustmanetal. (2001)
Noormets etal. (200 Ib)
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1 2002). This study also confirmed earlier observations of changes in trophic interactions
2 involving keystone tree species, as well as important insect pests and their natural enemies
3 (Table 9-26) (Percy et al., 2002; Holton et al., 2003; Awmack et al., 2003).
4
5 9.7.3 Landscape Condition
6 In the SAB framework (Figure 9-19), landscape condition is assessed using the areal
7 extent, composition of component landscape ecosystems or habitat types, and the pattern or
8 structure of component ecosystems or habitat types (including biocorridors). To date, no
9 publications exist on the impacts of O3 exposure on landscape condition. The effects of O3
10 exposure have only been reported at the community or stand level (see Biotic Conditions,
11 below). The following is a description of current discussions by land stewards and of how
12 difficult it will be to quantitatively assess the effect of O3 exposure on landscape condition.
13 Landscapes are identified and preserved, such as National Parks, Class I Wilderness Areas,
14 etc., so that they are protected from the effects of O3 exposure by law. Efforts to determine
15 whether landscapes have been affected by certain levels of exposure rely on valuation of
16 landscape and ecosystem components. Several different approaches of valuation have been
17 used, such as pathological (visible symptoms), biomass and allocation, and biogeochemical.
18 In the pathological approach, a "critical loads" concept is developed, with varying levels of
19 impact viewed as acceptable, interim targets, or as unacceptable. As an example, land managers
20 of Class I Wilderness Areas may consider a critical level acceptable if it resulted in no visible O3
21 symptoms to sensitive species. In concrete terms, sensitive species may respond to peak O3
22 exposures of 60 ppb (e.g., coneflower, Rudbeckia laciniata, in Great Smoky National Park;
23 Davison et al., 2003), and so the critical exposure level would be < 60 ppb for any hourly valued
24 during the growing season. An interim target would be that less than 5% of the sensitive plants
25 would have visible symptoms of < 15% of the leaf surface. An unacceptable level of O3
26 exposure would be any result more pronounced than the interim target. The advantage of the
27 foliar injury approach is that large crews with relatively simple training can assess individual
28 species within the landscape and "see" the effect of the oxidant exposure. There are several
29 disadvantages, however. Some species (e.g., white fir) exhibit no foliar injury, but do have shifts
30 in biomass allocation in response to oxidant exposure (Retzlaff et al., 2001). Other species have
31 shown significant decreases in foliar injury due to needle loss, retranslocation of nutrients to
January 2005 9-275 DRAFT-DO NOT QUOTE OR CITE
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Sources
Above-ground Processes
(-) Rubisco activity
(+) Lipid Peroxidation
(-) Stomatal Control
(-) Phloem Loading
(+) Senescence
Sinks
Above-ground Processes
(+) Antioxidant Synthesis
(+) Repair Processes
(+) Construction Costs
(+) Tissue Respiration
(-) Growth
(±) Reproduction
Storage Carbohydrates
(-) Pool Size
(-) Concentration
Root Storage
(-) Pool Size
(-) Concentration
Storage Carbohydrates
(-) Pool Size
(±) Concentration
Root Processes
(-) Growth
(±) Mycorrhizal
Colonization
(?) Turnover
Root Storage
(-) Pool Size
(-) Concentration
Figure 9-19. A conceptual diagram of processes and storage pools in sources and sinks
that are affected by ozone exposure. A plus (+) denotes an increase in
process rate or pool size, a minus (-) denotes a decrease in process rate or
pool size, and a plus-minus (+) denotes that both increases and decreases
have been reported in response to O3. Primary effects in the shoots (1°) are
distinguished from secondary effects in roots (2°) since the primary site of
ozone action occurs in the leaves (from Andersen, 2003).
1 remaining foliage, with subsequent increased photosynthetic rate (Beyers et al., 1992).
2 In addition, the development of foliar symptoms within a species is related to sunlight and
3 microclimate (Davison et al., 2003).
4 In the biomass approach, O3 exposure resulting in a measureable decline in biomass
5 (usually of a target, sensitive species) is used to evaluate landscape condition. The bulk of the
6 information available is from seedling responses to controlled chamber exposures, reviewed in
7 the previous section. Some information exists for species in natural environments, but teasing
8 out concurrent stressors and finding adequate "controls" may be intractable. For example, in a
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1 long-term gradient of O3 exposure, N deposition, and drought, the site with the highest O3
2 exposure had the greatest whole tree biomass (pole-sized trees) due to growth stimulation by N
3 deposition (Grulke and Balduman, 1999).
4 In the biogeochemical approach, changes in biogeochemical cycling are used to assess
5 landscape condition. Species sensitive to O3 exposure have well known responses to O3
6 exposure, including altered C allocation to below- and above-ground tissues, and altered rates of
7 leaf production, turnover, and decomposition. Changes in turnover rates of ephemeral tissues
8 (leaves, fine roots) also affect nutritional status of the remaining tissue. These shifts can affect
9 overall carbon and N loss from the ecosystem in terms of respired C, and leached aqueous
10 dissolved organic and inorganic C and N. Instability in C and N pools and dynamics can affect
11 landscape-level nutrient dynamics even without significant inputs of N deposition. The endpoint
12 assessment is based on changes in water quality from or in the landscape, correlated to a defined
13 oxidant exposure level. These approaches are linkable: visible injury at a particular level could
14 be related to reduction in photosynthate, which would reduce whole plant biomass (and carbon
15 dynamics). If O3 sensitive species are dominant within the landscape, then changes in C and N
16 dynamics over time would be expected to alter biogeochemical cycles. Examples of forest types
17 that contain geographically extensive, O3-sensitive species that could be used in assessing
18 landscape-level changes include ponderosa pine in the western United States, tulip poplar
19 (Liriodendron tulipiferd) or loblolly pine (Pinus taeda L.) in the eastern U.S. decidous forests,
20 and Norway spruce (Picea abies (L.) Karst) in the Carpathian Mountains of eastern Europe.
21 Water quantity may also be affected by O3 exposure at the landscape level. Moderately
22 high O3 exposure may affect the mechanism of stomatal opening (McAinsh et al., 2002),
23 resulting in sluggish stomatal opening and closing (Patterson and Rundel, 1989; Reich and
24 Lassoie, 1984). During moderately high O3 exposure in a drought year, canopy transpiration was
25 greater for tulip poplar than on adjacent days with lower O3 exposure. This could affect canopy
26 transpiration and landscape level water use (McLaughlin et al., 2004). Oxidant exposure (O3 and
27 NOX) may decrease the ability of exposed plants to close stomata at night (Grulke et al., 2004),
28 thus increasing water loss from the landscape. Ecosystem models should aid in interpreting
29 O3-exposure effects at the landscape level.
30
January 2005 9-277 DRAFT-DO NOT QUOTE OR CITE
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1 9.7.4 Biotic Condition
2 9.7.4.1 Ecosystems and Communities
3 The SAB framework described by Young and Sanzone (2000) (Figure 9-20) identifies
4 community extent, community composition, trophic structure, community dynamics, and
5 physical structure as EEA's for assessing ecosystem health.
6
7 COMMUNITY EXTENT
8 Ecosystem function is dependent on areal extent, constituent species composition, trophic
9 structure and its dynamics, and community physical structure. Genetic variation within species,
10 and the dynamics of the interactions that exist among different species and their biotic and
11 abiotic environment, are also involved (Agrawal and Agrawal, 2000). There are no reports of O3
12 exposure altering community distribution or extent.
13
14 COMMUNITY COMPOSITION
15 Significant changes in plant community composition resulting directly from O3 exposure
16 has been demonstrated in two forested areas: the mixed conifer forest of the San Bernardino
17 Mountains, CA and the mixed conifer forest of the Valley of Mexico near Mexico City. It is
18 also likely that community composition has changed in response to O3 exposure in the
19 coniferous forests of the Carpathian Mountains, but this has not yet been definitively shown.
20 The first forest communities shown to be affected by O3 were the Pinus ponderosa-
21 dominated stands of the San Bernardino Mountains in southern California (Miller, 1973). Miller
22 suggested that mixed forests of P. ponderosa, Pinus jeffreyi and Abies concolor were changing
23 to predominantly A. concolor because of the greater sensitivity of the pines to O3. Significantly
24 greater mortality of young mature trees (50 to 99 years old) occurred in sites that also showed
25 higher foliar injury relative to sites that showed slight foliar injury (McBride and Laven, 1999).
26 For P. ponderosa, 33% of the trees in the high foliar injury sites died versus 7% of the trees in
27 the low foliar injury sites over the decade-long census. In contrast, 24% of Abies concolor died
28 in high foliar injury sites, whereas no trees died in slight injury sites. The authors suggested that
29 certain age classes were especially sensitive to O3 exposure because they are emerging into the
30 canopy, where higher O3 concentrations are encountered. Future projections based on past
31 changes in community composition have been conducted for 2024 and 2074 (McBride and
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1 Laven 1999). In their projections, the population of Pinusponderosa nearly disappears in all
2 tree age classes, and the community is dominated by Quercus kelloggii in all tree age classes,
3 followed by Calocedrus decurrem and Pinus lambertiana by the year 2074. Their projections
4 do not account for potential changes in genetic structure of the more O3-sensitive species.
5 In the Valley of Mexico, the closed forest structure changed to a woodland from high
6 pollutant exposure (Miller et al., 2002). Cryptogamic community diversity also significantly
7 declined in response to prolonged, extreme O3 exposure (Zambrano et al., 2002. Together,
8 these two examples illustrate the potential for shifts in community composition in response to
9 O3 stress.
10
11 TROPHIC STRUCTURE
12 A bove-Ground Structural
13 One of the first reports of trophic level interactions in natural communities was the
14 O3-induced predisposition of Pinus ponderosa to attack by bark beetles (Stark et al., 1968; Cobb
15 et al., 1968; Stark and Cobb, 1969). Trees exposed to oxidant injury had lower resin production,
16 flow, and exudation pressure. Also, several attributes associated with tree defense against beetle
17 attack were compromised by oxidant exposure including sapwood and phloem moisture content
18 and phloem thickness (Pronos et al., 1999). Another trophic level has been implicated, in that O3
19 injured ponderosa pine had the same rate of bark beetle infection, but healthy trees had greater
20 numbers of bark beetle predators and parasitoids (Dahlsten et al., 1997). This suggests that O3
21 damage rendered the pines inhospitable for the natural enemies of the bark beetles. Similar
22 findings were presented by Percy et al. (2002) for aphids whose abundance was increased in
23 young Populus tremuloides stands exposed to elevated O3. In that study, the levels of natural
24 enemies of alphids (ladybirds, lacewings, spiders and parasitoids) were significantly decreased
25 under elevated O3.
26
27 Below-Ground
28 Processing of plant-derived carbon compounds by soil organisms comprising the soil food
29 web is a fundamental property of a functional and stable below-ground ecosystem (de Ruiter
30 et al., 1998; Wolters, 1998). Soil food web organisms are responsible for recycling nutrients and
31 for development of soil properties such as porosity, aggregate structure, water-holding capacity
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1 and cation exchange capacity. A shift in food web species diversity or functional complexity in
2 response to O3 stress may alter ecosystem processes.
3 Evidence that soil organisms are affected by O3 indicates the potential for changes in soil
4 food web structure and function. Since O3 does not penetrate the soil beyond a few centimeters,
5 the proposed mechanism by which O3 alters soil biota is through a change in carbon input to
6 soils (Andersen, 2003). Ozone can alter C inputs to soil and hence soil processes through four
7 different pathways: (1) leaf-litter quality and quantity (see Material Cycling, below), (2) carbon
8 allocation to roots (see Physiological Status, below), (3) interactions among root symbionts, and
9 (4) rhizodeposition. The complex nature of the effects of O3 on trophic interactions and food
10 webs calls for additional basic research and modeling.
11 There have been no comprehensive studies on the effects of O3 on structural components of
12 soil food webs, however studies have shown that O3 affects free-living soil organisms of food
13 webs. In the few cases where soil microbial communities have been examined, O3 has led to
14 changes in bacterial and fungal biomass, and in some cases changes in soil enzyme activity.
15 Phillips et al. (2002) examined the effects of elevated CO2 and O3 on C flow through
16 heterotrophic microbial communities in soils collected from a FACE study in Wisconsin. Ozone
17 decreased abundance of fungal phospholipid fatty acids in aspen and birch-aspen plots but had
18 few other direct effects on measured soil parameters. The greatest effect of O3 was to eliminate
19 significant increases in microbial respiration resulting from elevated CO2, suggesting an
20 important role for O3 in altering C flow through soils. Shafer (1988) found that O3 tended to
21 increase the number of fungal propagules and bacteria exhibiting phosphatase activity in the
22 rhizosphere of sorghum. Ozone in combination with simulated acid rain stimulated soil
23 arylsufatase activity (Reddy et al., 1991). The response was observed at low concentrations, but
24 was reversed at high concentrations, suggesting a threshold level of O3, possibly involving
25 different mechanisms. Ozone significantly decreased soil microbial biomass in the fall after one
26 season of exposure in a wheat (Triticum aestivuni) and soybean (Glycine max) system (Islam
27 et al., 2000). Other studies have shown shifts in microbial and fungal biomass in response to O3
28 stress, but responses were variable (Scagel and Andersen, 1997; Yoshida et al., 2001).
29 Decreased allocation to roots associated with O3 exposure alters N fixation in legumes and
30 actinorrhizal species. Ozone exposure was found to decrease nodulation in a number of species
31 (Manning et al., 1971; Tingey and Blum, 1973). In alder (Alnus sermlata), host root cells of
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1 nodules showed cytoplasmic breakdown and lacked organelles when seedlings were exposed to
2 27 d of O3 (Greitner and Winner, 1989).
3 Ozone has been shown to affect mycorrhizal colonization (Ho and Trappe, 1981; McCool
4 et al., 1982; Simmons and Kelly 1989; Adams and O'Neill, 1991; Edwards and Kelly 1992;
5 Smith and Read, 1997). Although short-term in nature, several studies have found enhanced
6 mycorrhizal short-root formation under O3 stress. White pine (Pinus strobus) (Stroo et al.,
7 1988), Norway spruce (Rantanen et al., 1994), Northern red oak (Quercus rubra) (Reich et al,
8 1985), Douglas-fir (Pseudotsuga menziesii) (Gorissen et al., 1991), European silver-fir (Abies
9 alba) (Wollmer and Kottke, 1990), and Scots pine (P. sylvestris) (Kasurinen et al., 1999) all
10 showed some increase in mycorrhizal presence when exposed to O3. Others have shown
11 minimal or no effects of O3 on mycorrhizas (Mahoney et al., 1985; Meier et al., 1990;
12 Kainulainen et al., 2000). Stroo et al. (1988) found that percent infection increased from 0.02 to
13 0.06 ppm O3, then decreased from 0.06 to 0.14 ppm; the total number of short roots were
14 unaffected, however. In cases where stimulation was observed, the response was often noted
15 shortly after initiation of exposure, often at relatively low concentrations. Good examples of this
16 transitory response can be found in results with Norway spruce and Scots pine (Ratanen et al.,
17 1994; Kasurinen et al., 1999). In these studies, O3 increased mycorrhizal short roots initially but
18 differences were not evident by the end of the experiment.
19 There is evidence which suggests that decreased below-ground allocation associated with
20 O3 stress alters mycorrhizal host-symbiont compatibility. Edwards and Kelly (1992) found a
21 shift in fungal morphotypes present on loblolly pine roots, even though the number of
22 mycorrhizal short roots per gram fine root was not significantly affected by O3. Qiu et al. (1993)
23 found increased numbers of morphotypes present on O3-sensitive loblolly pine seedlings exposed
24 to O3. Roth and Fahey (1998) found an interaction between O3 and acid precipitation treatments
25 on the composition of fungi forming ectomycorrhizae on red spruce saplings, possibly driven by
26 nutrient availability. Carbohydrate requirements vary among fungal species (Bidartondo et al.,
27 2001), and O3 may affect species composition by altering carbohydrate availability in roots.
28 A shift in species dominance could lead to a change in successional patterns of mycorrhizal
29 communities.
30 In the few studies that examined root exudation in response to O3 exposure, O3 was found
31 to alter the quantity and quality of root exudates. McCool and Menge (1983) found a significant
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1 decrease in exudation of amino acids in tomato (L. esculentum) exposed to 300 ppb O3. McCrady
2 and Andersen (2000) observed increased root exudation in nonmycorrhizal wheat (Triticum
3 aestivum). No apparent change in root exudation was found in labeling studies of ECM
4 ponderosa pine (Andersen and Rygiewicz, 1995). Inconsistency in the literature probably results
5 from species differences and experimental protocols, however these examples illustrate the
6 potential effects of O3 on rhizosphere carbon flux.
7 Decreased carbon allocation to roots of O3-exposed plants may reduce root longevity and
8 accelerate root turnover, increasing rhizodeposition of C and N. Fine root turnover decreased in
9 mature northern red oak exposed to elevated O3 (seasonal exposure ranging from 152 to
10 189 ppm-h), whereas seedlings did not show any reduction in turnover (Kelting et al., 1995).
11 King et al. (2001) found a trend toward decreased live root biomass and increased dead root
12 biomass in aspen exposed to O3 in a FACE study, suggesting possible changes in both
13 product! on and 1 ongevity.
14 Other studies also suggest that O3 alters C flux to soils, resulting in changes in CO2 efflux
15 from soils. Both root respiration and soil CO2 efflux decreased from loblolly pine seedlings
16 exposed to O3 (Edwards, 1991). Soil CO2 efflux increased in response to O3 in ponderosa pine
17 seedlings (Andersen and Scagel, 1997; Scagel and Andersen, 1997). No direct assessments of
18 hyphal growth and turnover in response to O3 stress have been conducted. Ozone decreased C
19 allocation to extrametrical hyphae of a ponderosa pine mycorrhiza, which might be expected to
20 decrease growth and increase hyphal turnover (Andersen and Rygiewicz, 1995).
21
22 COMMUNITY DYNAMICS, PHYSICAL STRUCTURE
23 One of the best-documented examples of change in long-term forest community dynamics
24 of dominant overstory trees occurred in the San Bernardino Mountains between 1968 and 1974
25 (reported in Miller, 1973; Miller et al., 1989). Plots were recently re-inventoried -25 years after
26 establishment (Arbaugh et al., 2003). Of the six codominant canopy species, white fir showed
27 the greatest change, increasing in both numbers and bole growth for a 286% change in basal
28 area/ha in the San Bernardino Mountains. Sugar pine basal area also increased significantly (by
29 334%), but this species represents only a small portion (1%) of the total basal area of the forest
30 sampled. The most sensitive species (Miller et al., 1983), ponderosa and Jeffrey pine, had the
31 lowest increase in basal area/ha (76 and 62%, respectively). These two species represented 72%
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1 of the basal area/ha of all stands inventoried. Ponderosa pine had the greatest mortality rate of
2 all canopy species inventoried (46%), followed by white fir and black oak (35% and 33%),
3 Jeffrey pine (29%) and incense cedar and sugar pine (both 7%). In moist sites (at the western
4 end of the San Bernardino Mountains), there was significant recruitment of incense cedar, white
5 fir, and sugar pine. Only one study directly attributed tree mortality to O3 exposure: it
6 accounted for 7% of mortality in the Sierra and Sequoia National Forests (Carroll et al., 2003).
7 Species diversity in the understory can be quite large, making studies of O3 effects on
8 understory community dynamics very challenging. However, there have been some attempts to
9 quantify understory responses, ranging from describing relative sensitivity to their visible
10 symptoms (Treshow and Stewart, 1973; Temple, 1999) to very complex measures of community
11 structure and composition (Westman, 1979, 1981). The lowest percentage cover and lowest
12 species diversity in California coastal sage scrub was correlated with the highest O3 exposures as
13 estimated by extrapolation from the closest air monitoring stations (Westman 1979). The
14 understory also has the potential to influence responses to O3 of dominant keystone species, as
15 has been shown in controlled experiments with bothponderosapine (Andersen et al., 2001) and
16 loblloly pines (Barbo et al., 2002). Barbo et al., (1998) exposed an early successional forest
17 community to ambient air, charcoal-filtered air, non-filtered air, and 2x ambient in the
18 Shenandoah National Park. They found changes in species performance, canopy structure,
19 species richness and diversity index consistent with the view that O3 can induce a shift in
20 vegetation dominance and community structure.
21 There have been few studies evaluating the effect of O3 exposure on the physical structure
22 of natural ecosystems. Despite an extensive array of allometric equations for conifers in the
23 west on United States (Ter-Mikaelian and Korzukhin, 1997), none appear to predict individual
24 tree shape in a site of moderate O3 exposure, suggesting that O3 may effect allometry (Grulke
25 et al., 2003a). Canopy structural changes are also implied by the measure of canopy
26 transparency used in the FHM assessment. The loss of epiphytic lichens within the canopy is a
27 clear example of plant community structural change occurring along an O3 gradient (Nash and
28 Sigal, 1999; Zambrano et al., 2002).
29 As of yet, there have been no comprehensive studies on the effects of O3 on structural or
30 functional components below-ground (Andersen, 2003). Phillips et al. (2002) found evidence
31 for changes in the bacterial and fungal biomass below Populus tremuloides and
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1 P. tremuloideslBetulapapyri/era stands exposed to elevated O3. Subsequent study showed that
2 O3 exposure decreased cellobiohydrolase activity in the soil microorganisms, driving the change
3 in the microbial community (Larson et al., 2002).
4
5 9.7.4.2 Species and Populations
6 Ozone can affect species and populations of species comprising ecosystems through
7 changes in population size, genetic diversity, population structure and/or dynamics, and habitat
8 suitability (Young and Sanzone, 2002). For example, if individuals of a species are lost due to
9 O3 exposure, population size declines. Often very young (e.g., conifer seedlings, see Section
10 9.7.4.3 below) and old individuals differ in their sensitivity, so that population structure also will
11 be altered by O3 exposure. If resource allocation to reproductive output is altered by O3
12 exposure, population dynamics will be altered. Communities dominated by O3 sensitive species
13 in the canopy or understory may be altered sufficiently for the habitat to become unsuitable for
14 other species. Genetic selection acts on the individual plant, which represents a certain
15 proportion of the populations' genetic variation. If an O3 sensitive individual succumbs through
16 multiple stresses, including O3 stress, the genetic variation represented in the population
17 generally declines, unless sensitive individuals have low inherent genetic variability (e.g.,
18 Staszak et al., 2004).
19 While the concept of natural selection induced by O3 exposure and related changes in
20 natural plant communities has been around for a long time (Dunn, 1959; Miller et al., 1972), the
21 concept of evolution of O3 tolerance is still not widely accepted. The unequivocal demonstration
22 that considerable genetic variation in O3 resistance exists within and between plant populations,
23 and that ambient levels of O3 may differentially affect fitness-related traits (i.e., growth, survival,
24 and fecundity), suggests that O3 may potentially drive the natural selection for resistant
25 individuals. Dunn (1959) presented circumstantial evidence that ambient O3 in the Los Angeles
26 area was high enough to drive the selection of O3 resistant Lupinus bicolor genotypes. Since
27 Dunn's (1959) research on O3-induced population changes, researchers have demonstrated
28 differences in O3 tolerance among other plant populations. In the devastating forest decline
29 southeast of Mexico City, the remaining trees (primarily Abies religiosa in the "cemetery
30 forests") appear to be less affected by foliar injury than trees that were lost (Alvarado-Rosales
31 and Hernandez-Tejeda, 2002), despite continued, high O3 exposures (Bravo-Alvarez and
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1 Torres-Jardon, 2002). However, even the most convincing work in this field (Berrang et al.,
2 1986, 1989, 1991), whhPopulus tremuloides, where a strong correlation between visible foliar
3 injury after O3 exposure and maximum O3 concentration at the origin of the population was
4 shown (Berrang et al., 1991), a change in gene frequency at any one site over time has not yet
5 been demonstrated (Bell et al., 1991; Reiling and Davison, 1992b). Furthermore, the selection
6 intensity of O3 has been questioned (Bell et al., 1991; Taylor and Pitelka, 1992; Taylor et al.,
7 1994) and the emergence of O3 exposure since the 1950's as an environmental stressor may not
8 have been long enough to affect tree populations with long generation times (Barrett and Bush,
9 1991).
10 The loss of O3-sensitive individuals results in natural selection favoring O3-tolerant species
11 (Bradshaw and McNeilly, 1991). Increased levels of mortality of O3-sensitive individuals have
12 occurred for Pinus jeffreyi and P. ponderosa exposed to ambient O3 along the western slope of
13 the Sierra Nevadas (Peterson et al., 1987; Miller et al., 1998), for Pinus strobus exposed to
14 ambient O3 in southern Wisconsin (Karnosky, 1981), and for P. ponderosa in the San Bernardino
15 Mountains (Carroll et al., 2003). In these examples, individuals that consistently had greater
16 foliar injury and lower needle retention were lost in repeated surveys. Ozone-induced loss of all
17 individuals except the most tolerant and breeding among the surviving individuals to yield more
18 more tolerant populations has not yet been demonstrated for plants exposed to O3, except for the
19 relatively short term (2 years) adaptation exhibited in Trifolium repens (Heagle et al., 1991) and
20 Plantago major (Davison and Reiling, 1995). Heagle et al. (1991) were able to show the
21 adaptation of a Trifolium repens population to elevated O3 in just two growing seasons.
22 Similarly, Davison and Reiling (1995) compared the O3 resistance of P. major populations
23 grown from seed collected from the same sites over a period of increasing O3. The two
24 independent populations studied exhibited increased O3 resistance, consistent with the idea of
25 selection for O3 tolerance. Using random amplified polymorphic DNA primers, this team also
26 showed that the later populations are subsets of the earlier ones, consistent with in-situ evolution
27 rather than with catastrophic loss and replacement of the populations (Wolff et al., 2000). The
28 problem is determining whether spatial patterns in O3 resistance and changes in time are casually
29 related to O3, because there were very strong correlations with other factors (Reiling and
30 Davison, 1992a; Davison and Barnes, 1998). The potential for evolution of O3 resistance has
31 been clearly demonstrated by Whitfield et al. (1997) in their study of O3 selection of common
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1 plantain (Plantago major L), where they showed that within a matter of a few generations, it
2 was possible to increase O3 resistance in an initially O3-sensitive population. Wild radish
3 (Raphanus sativus L.) developed O3 resistance after only one generation of exposure to O3
4 (Gillespie and Winner, 1989).
5 A third independent line of research suggesting O3 may be affecting the genetic diversity of
6 wild plant populations was presented by Paludan-Miiller et al. (1999) who showed that northwest
7 European provenances of European beech (Fagus sylvatica L.) were more sensitive to O3 than
8 were southeast European provenances which had experienced higher O3 levels. Recent research
9 on the genetic structure of 50-year-oldponderosapines in the San Bernardino Mountains
10 suggests that distinct differences in frequency of some alleles and genotypes occurred, with the
11 O3-tolerant trees being more heterozygous (Staszak et al., 2004). While both of these studies
12 were only correlational, they are consistent with previous studies of this type suggesting
13 O3-induced population changes. Again, other environmental stressors besides O3 exposure could
14 have been involved in effecting change within these populations.
15 Natural selection for O3 tolerance can also be facilitated by reductions in fitness related to
16 lower seed yields of O3-sensitive species or individuals. The impacts of O3 on reproductive
17 development, recently reviewed by Black et al. (2000) can occur by influencing: (1) age of
18 flowering, particularly in long-lived trees that often have long juvenile periods of early growth
19 without flower and seed production; (2) flower bud initiation and development; (3) pollen
20 germination and pollen tube growth; and (4) seed, fruit, or cone yields and seed quality
21 (Table 9-24). In addition, vegetatively propagated species can have lower numbers of
22 propagules under elevated O3 conditions (Table 9-24).
23 Several studies suggest that reproductive structures are clearly sensitive to O3 and that O3
24 can affect fitness of plants by affecting either the sporophytic or gametic generations. Decreased
25 numbers of flower spikes and seed capsules per plant were found for plantain growing under
26 elevated O3 (Reiling and Davison, 1992b; Pearson et al., 1996; Lyons and Barnes, 1998).
27 Similar responses were seen for Brassica campestris L. plants exposed to a single dose of
28 100 ppb O3 for 6 h. Stewart et al. (1996) and Bosac et al. (1998) reported an increase of flower
29 bud abortion for oilseed rape (Brassica napus L.) similarly exposed to a short duration of
30 elevated O3. Floral initiation period can be delayed in O3-sensitive plants, as was described for
31 dogbane (Apocynum androsaemifolium) grown under ambient O3 in the eastern United States.
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1 In one of the few comparisons of whole plant O3 sensitivities with that of male gametophytes,
2 Hormaza et al. (1996) found a high correlation of relative O3 sensitivity of pollen tube elongation
3 with that of O3 effects on net photosynthesis and relative growth rates for 6 species of fruit trees.
4 Clearly, the concept of O3 -induced genetic change is an area that needs additional research
5 attention. Repeated collections over time from wild populations receiving high O3 exposures to
6 examine population responses and relative sensitivity changes, the sampling of genetic diversity
7 along known O3 gradients, and the use of modern biotechnological approaches to characterize
8 and quantify genetic diversity are useful approaches to test for O3-induced impacts on diversity
9 in natural ecosystems.
10
11 9.7.4.3 Organism Condition
12 PHYSIOLOGICAL STATUS
13 The generalized effects of O3 exposure on plants are well known, and have been reviewed
14 from several viewpoints over the last decade (De Kok and Tausz, 2001; Heath and Taylor, 1997;
15 Pell et al., 1997; Schraudner et al., 1997; Matyssek et al., 1995; Darrall 1989; Reich, 1987; U.S.
16 Environmental Protection Agency, 1996). The topic of individual species response and
17 modification of response by other factors has been addressed thoroughly in Sections 9.4.3 and
18 9.4.4 of this chapter. Here, physiological changes in response to O3 that have been hypothesized
19 to lead to changes in ecosystem structure or function are highlighted.
20
21 Above-Ground Responses
22 The first critical step leading to O3 response is uptake and movement of O3 by the leaves,
23 leading to changes in C and nutrient relations that are thought to alter plant growth and
24 competiveness (see Section 9.3). Ozone enters leaves through stomates, reacts with cell walls or
25 membranes, and starts a series of adverse reactions. Cuticular uptake of O3 is believed to be
26 negligible (Kerstiens and Lendzian, 1989; Coe et al., 1995). Once inside the leaf, O3 and its
27 byproducts lead to membrane disruption, chlorophyll breakdown, and decreased Rubisco levels
28 (Schweizer and Arndt, 1990). In turn, photosynthesis is decreased, as is stomatal conductance
29 (Weber et al., 1993). Also, O3 often leads to increased maintenance respiration, decreased foliar
30 nutrient content, and imbalances in tissue nutrient content and retention. When photosynthetic
31 pigments have been damaged, the pigment must be fully broken down (and/or new N and Mg
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1 must be taken up and transported to the leaf) for the pigment to be regenerated (Bjorkman and
2 Demmig-Adams, 1995). Ozone exposure alters within-plant priorities for resources: less C is
3 available for allocation to roots and spring regrowth, and less foliar biomass is retained. At the
4 whole-organism level, O3 exposure decreases root mass (Grulke et al., 1998) and radial bole
5 growth (Peterson et al., 1991; Muzika et al., 2002) with little impact on height growth. Visible
6 symptoms of O3 injury vary between species and genotypes but often include upper leaf surface
7 stipple, chlorotic mottle, or large bifacial blotches of necrotic tissue. Premature senescence is
8 typical of almost all O3-induced foliar damage. All of these changes can alter the plant's ability
9 to function in a broader ecosystem context.
10 The underlying mechanisms of O3-injury response in conifers, broadleaf deciduous trees,
11 and herbaceous species are assumed to be similar. However, several differences in long-lived
12 species are important at the ecosystem level. Most of the research on O3 effects has been
13 conducted on herbaceous species (i.e., crops). Although a number of native herbaceous species
14 have been identified as O3-sensitive. there are no published physiological studies on the effect of
15 O3 exposure on herbaceous or shrub species in situ. In natural ecosystems, the majority of
16 species are not annuals, unless the system is highly disturbed. Nonetheless, response of crop
17 species to elevated O3 may be used as an analog for native annual response: phenological
18 staging is accelerated (soybeans; Booker et al., 2004), thus "avoiding" additional O3 exposure.
19 Conifers have roughly half the stomatal conductance of deciduous broadleaf trees (Reich,
20 1987), leading to proportionally less O3 uptake at the same O3 exposure level. Yet, except for
21 species of larch, individual conifer needles are longer-lived and active over a greater portion of
22 the year. Therefore, needle longevity can also work against the tree by increasing cumulative O3
23 exposure and exposure to other stressors. Increased needle longevity is not always a
24 disadvantage, for example, conifers are physiologically active in early spring and late fall, during
25 times of lower oxidant concentrations. These periods can contribute significantly to a net
26 positive annual C balance and from the standpoint of nutrient storage, are important in reparation
27 responses to pollutants. Patterson and Rundel (1995) reported that Jeffrey pine had significant
28 stomatal opening (one third that of a typical summer day) in mid-winter with snow on the
29 ground. At least pole-sized and larger trees can mitigate reductions in C acquisition due to
30 oxidant exposure in the summer with C assimilation on favorable days in the winter. The
31 interaction of environmental factors, plant phenology (the timing of growth events; birth and
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1 mortality of plant parts), physiological status (nutritional or moisture status; dormant or active
2 growth within the year), and tree age (interannual differences in resource acquisition and
3 allocation) all contribute to the complexity of long-lived species (and hence ecosystem) response
4 to O3 exposure.
5 One widely observed response to O3 exposure is premature leaf loss. As noted above,
6 premature leaf loss may reduce O3 uptake during high-O3 years, but it has several negative
7 consequences. Early leaf loss results in reduced C uptake through photosynthesis. Premature
8 needle loss also results in less N retranslocation compared to normally senescing leaves, which
9 reduces whole plant N balance (Fenn and Dunn, 1989) and carbohydrate availability for
10 overwinter storage (Grulke et al., 2001). Because of such effects accumulated over several years
11 of O3 exposure, subsequent-year C and N reserves can be affected (Andersen et al., 1991).
12 Conversely, a series of drought years can decrease O3 uptake, as well as reduce C and
13 nutrient acquisition, altering resource allocation to defenses (e.g., antioxidants) (Grulke et al.,
14 2003b) (or resins) against insect infestation, rendering the tree more susceptible to O3 injury.
15 Conifers have thicker cuticles than either broadleaf deciduous or herbaceous species. Continued
16 O3 exposure may compromise cuticular integrity (Percy et al., 1994). Once cuticular integrity is
17 breached, individual leaves (needles) are likely to be excised, thus contributing generally to
18 defoliation and reduced C acquisition.
19 With the exception of the extensive research conducted on mature tree response to O3
20 exposure in California forests (Peterson et al., 1987, 1991, 1995, Arbaugh et al., 1998; Grulke
21 1999; Grulke and Balduman 1999; Grulke et al., 1996, 1998, 2001, 2002b, 2003a,b, 2004;
22 Weiser et al., 2002), the vast majority of studies of O3 effects on forest trees have been
23 conducted on young seedlings (Chappelka and Samuelson, 1998) and little is known about
24 acclimation to O3 (Skarby et al., 1998). Chamber exposure studies can be used to document
25 foliar symptoms and develop response variables for the whole plant. These response variables
26 can then be field tested on mature trees using correlative analyses (e.g., Grulke and Lee, 1997;
27 Grulke et al., 2003b). Without the initial work in chamber exposure studies, field responses to
28 O3 exposure would be difficult to verify and distinguish from other concurrent stressors.
29 Predicting mature tree responses to O3 solely from seedling response studies is complex,
30 because seedlings or saplings do not necessarily respond to O3 in the same way as mature trees
31 (Norby et al., 1999; Karnosky, 2003). Ozone has been found to have stronger effects on leaf
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1 function in younger rather than older trees (Kolb and Matyssek, 2001). Each component
2 physiological attribute "matures" at a different rate. Gas exchange patterns differ between
3 seedlings and mature trees. For example, leaf respiration of juvenile ponderosa pine was greater
4 than that of mature trees (Momen et al., 1996). In the conifers tested, the highest gas exchange
5 rates (and by inference stomatal uptake of O3) are found in seedlings (e.g., in scions of red
6 spruce, Rebbeck and Jensen, 1993; giant sequoia, Grulke et al., 1994; ponderosa pine, Grulke
7 and Retzlaff, 2001; and Norway spruce, Wieser et al., 2002). Patterns of biomass (Grulke and
8 Balduman, 1999) and carbohydrate allocation (Grulke et al., 2001) differs between immature and
9 mature trees. Pole-sized trees had greater reduction in root, foliar, and bole carbohydrate
10 concentrations than did old growth trees. Antioxidant defenses vary with both tree age and
11 needle age (Tegisher et al., 2002). Based on all attributes measured in both ponderosa pine and
12 giant sequoia, the youngest tree age considered representative of mature trees was 20 years old
13 (Grulke et al., 1996; Grulke and Retzlaff, 2001). In some broadleaf deciduous tree species,
14 seedlings are more conservative, and mature trees have greater gas exchange rates, as is the case
15 for Quercus robur (Edwards et al., 1994; Kelting et al., 1995; Samuelson and Kelly, 1996, 1997)
16 and Fagus sylvatica (Braun et al., 1999). In another broadleaf deciduous tree species (cherry,
17 Prunus serotina), gas-exchange rates of seedlings were faster, but total O3 flux to leaves of
18 seedlings was lower than that of mature trees due to differences in leaf ontogeny (Frederickson
19 etal., 1995, 1996).
20 Nitrogen deposition modifies the effects of oxidant exposure through several offsetting
21 physiological mechanisms (see Section 9.4.4). Nitrogen deposition, in wet or dry particulate
22 form, ultimately increases site fertility, but increased soil N availability decreases C allocation to
23 roots, further exacerbating the effects of O3 exposure on roots (Grulke et al., 1998). Increased N
24 availability also increases foliage turnover: fewer needle age classes are retained (Gower et al.,
25 1993). Therefore, the combination of both increased N and O3 exposure increases foliar
26 turnover. Finally, N deposition and increased plant N nutrition can increase stomatal
27 conductance, leading to increased O3 uptake. Alternatively, increased N counteracts the effect of
28 O3 on photosynthesis by increasing photosynthetic pigments and enzymes. Nitrogen deposition
29 may mitigate the degree of foliar injury from oxidant pollution via higher available N for
30 reparation of photosynthetic pigments. Nitrogen amendments also modify the antioxidant
31 defense system in complex ways (Polle, 1998).
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1 Attributes of O3 injury to trees (foliar injury, needle retention, and canopy transparency), as
2 well as presence of pathogens and insect infestation, are routinely inventoried in established
3 sample plots distributed on Federal lands across the United States (Forest Health Protection,
4 USDA Forest Service). Foliar injury to several widespread, herbaceous species nationally
5 recognized as sensitive (bioindicators) is also assessed (NFS, 2003). This assessment is part of a
6 larger assessment of forest tree growth and dynamics (the Forest Inventory and Analysis
7 Program; Smith, 2002; Smith et al., 2003). Risk of O3 injury is then estimated for the dominant
8 forest tree species in the sample plots. For example, 12% of sampled black cherry (Primus
9 serotina), 15% of loblolly pine (Pinus taeda), and 24% of sweetgum (Liquidambar styraciflud)
10 were found to be in the highest risk category in the northeast and mid-Atlantic states (Coulston
11 et al., 2003). In the Carpathian Mountains, 12 to 13% of all trees (broadleaf and coniferous)
12 have greater than 26% crown defoliation (Badea et al., 2002). In general, broadleaves (primarily
13 beech) trees were less affected (8 to 45%) than spruce (up to 37%) and fir (up to 50%)
14 (Grodzinska et al., 2002). Ozone injury was directly correlated with cumulative O3 exposure in
15 the Sierra Nevada Mountains (Arbaugh et al., 1998); with the best correlation being found across
16 sites where > 90% of the trees had O3 injury. Although direct links of visible foliar symptoms
17 induced by O3 to adverse effects on biomass are not always found, visible foliar symptoms have
18 been linked to decreased vegetative growth (Peterson et al., 1987; Karnosky et al., 1996; Somers
19 et al., 1998), as well as reproductive function (Black et al., 2000; Chappelka, 2002).
20 Foliar O3 injury has also been associated with adverse effects on competitive ability and
21 survival in forest communities (Karnosky, 1981; McDonald et al., 2002; Karnosky et al., 2003b).
22 Competition can alter organism condition and affect susceptibility to O3. Ponderosa pine
23 seedlings were more susceptible to O3, as determined by decreased plant biomass, when grown
24 in competition with blue wild-rye grass (Andersen et al. 2001). Similarly, the magnitude of O3
25 effects on height and diameter growth depended on the competitive status ofPopulus
26 tremuloides trees (McDonald et al., 2002). These studies show the importance of including
27 competition as a concurrent stressor in assessing whole plant responses to O3. The age of the
28 community ("time since disturbance") may also affect the ability of individuals to effectively
29 respond to O3. Unfortunately, the vast majority of O3 studies have been conducted on
30 open-grown plants, often grown in pots where competition is absent both above and below
31 ground.
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1 Clearly, age-dependent O3 responsiveness and juvenile-mature correlations remain
2 important research questions in attempting to scale up to ecosystem level responses. Patterns of
3 allocation between root, stem, and leaf differ between immature and mature trees. Tree
4 architecture varies with tree age, and leaf area distribution in space and time may change in
5 response to elevated O3. All of these factors influence gas exchange in the canopy.
6 Furthermore, there may be few generalities that can be made from seedling to mature tree
7 response to O3 within a plant functional group (Norby et al., 1999; Karnosky, 2003).
8 Consequently, modeling ecosystem response is limited to either dealing with mono-specific
9 plantations or assigning average responses to a mix of species.
10
11 Below-Ground Responses
12 The effect of O3 on the soil ecosystem is thought to occur through physiological changes in
13 the root and interactions with soil organisms (Andersen, 2003). Comparatively little is known
14 about how changes in root growth and metabolism are translated through the soil food web,
15 resulting in changes in soil and hence, ecosystem processes. An overview of physiological
16 changes likely to lead to changes at the ecosystem level is provided below.
17 Ozone stress decreases carbon allocation to roots (Manning et al., 1971; McLaughlin and
18 McConathy, 1983; McCool and Menge, 1983; Cooley and Manning, 1987; Gorissen and van
19 Veen, 1988; Spence et al., 1990; Gorissen et al., 1994; Rennenberg et al., 1996; U.S.
20 Environmental Protection Agency, 1996). Since roots are often dependent on current
21 photosynthate for their structural development (van den Driessche, 1978; Ritchie and Dunlap,
22 1980; Marshall and Waring, 1985; van den Driessche, 1991), C-limiting stresses such as O3 can
23 have rapid and significant effects on root growth. In many cases, decreased allocation to roots in
24 response to O3 occurs quickly, with reductions in root growth occurring within one growing
25 season (Gorissen and van Veen, 1988; Spence et al., 1990; Gorissen et al., 1991; Andersen and
26 Rygiewicz, 1991, 1995; U.S. Environmental Protection Agency, 1996). Decreased C allocation
27 below ground is often associated with decreased root-shoot ratio, but observed responses in
28 root-shoot ratio are highly variable owing to several factors including intra- and interspecies
29 variation, culture conditions, and ontogenetic drift (Reich, 2002). Root-shoot ratio is a
30 point-in-time measurement that does not include C lost to exudation, respiration or turnover.
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1 Therefore, biomass and ratios of biomass (such as root-shoot ratio) do not necessarily reveal
2 physiological changes in response to O3 stress.
3 Decreased C acquisition leads to reduced carbohydrate levels and storage pools in
4 O3-exposed plants (Tingey et al., 1976; Ito et al., 1985; Cooley and Manning, 1987; Rebbeck
5 et al., 1988; Gorissen et al., 1994; Andersen et al., 1997; McLaughlin et al., 1982). Although it
6 is difficult to quantify changes in the field, Grulke et al. (1998) found decreased medium and
7 fine root biomass with increased pollutant load across an O3 gradient in southern California.
8 Coarse and fine root starch concentrations also were lowest in mature trees at the most polluted
9 site (Grulke et al., 2001). The effects of O3 could not be completely separated from other known
10 stresses across the pollutant gradient, but it appeared that O3 was an important factor in the
11 patterns observed.
12 Decreased storage pools can lead to carry-over effects on root growth that are compounded
13 over time. Decreased carbohydrate storage pools were associated with decreased root growth
14 during the spring following exposure to O3, even in the absence of additional O3 exposure
15 (Andersen et al., 1991, 1997). Decreased spring root growth was attributed to decreased stored
16 C reserves as well as to premature loss of older foliage age classes during the previous fall.
17 Aside from the loss of photosynthetic surface area associated with premature senescence, early
18 loss of foliage in the fall occurs when allocation to roots is at a maximum in many species
19 (Kozlowski and Pallardy, 1997). Older needle age classes preferentially allocate photosynthate
20 basipetally to stems and roots (Rangnekar et al., 1969; Gordon and Larson, 1970), and the loss of
21 older needles in the fall during allocation to root growth and storage, and in the spring during
22 periods of root growth, preferentially impacts roots and root processes.
23 Ozone has also been shown to affect root metabolism as evidenced by changes in root
24 respiration. Edwards (1991) found decreased root and soil CO2 efflux during a 2-year exposure
25 of loblolly pine to O3. Fine root respiration increased in mature red oak exposed to O3, while
26 total soil CO2 efflux increased in the spring and decreased in the summer and fall (Kelting et al.,
27 1995). The authors attributed increased root respiration to increased nutrient uptake in support
28 of increased demands in the shoot. Ozone decreased root system respiration in aspen after 12
29 weeks of exposure, but the decrease was closely associated with decreased root biomass and
30 probably not metabolic processes (Coleman et al., 1996). Whether other metabolic shifts occur
31 in the roots of plants exposed to O3 needs to be examined.
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1 Measurable effects on roots may occur before effects on shoots are observed because
2 shoots have immediate access to C for repair and compensation whereas roots must compete
3 with shoots for C. Mortensen (1998) found decreased root but not shoot growth in Betula
4 pubescens at O3 exposures of 42 nMol mol-L (applied 12 h d"1), whereas both root and shoot
5 growth were reduced at higher exposures. Chromosomal aberrations were found in root tips of
6 Norway spruce exposed to O3, even in the absence of biochemical changes in needles (Wonisch
7 et al., 1998, 1999). Using relatively high O3 concentrations (0.15 ppm O3 6 h d"1), Hofstra et al.
8 (1981) found metabolic changes in Phaseolus vulgaris root tips prior to the development of leaf
9 injury. Morphological changes in root tips occurred within 2 to 3 days, and metabolism declined
10 within 4 to 5 days of initiation of O3 exposure.
11 Feedback signals from roots can influence the degree of O3 response. Stolzy et al. (1964)
12 exposed tomato roots (Lycopersicon esculentum) to periods of anerobic conditions and followed
13 a change in leaf susceptibility to O3. An exposure of roots to low oxygen conditions for 3 h did
14 not alter photosynthesis, but foliar damage was decreased when the roots were subsequently
15 exposed to O3. In this case, a signal originating in the root appeared to alter leaf sensitivity to
16 O3, the signal possibly being hydraulic in nature and leading to decreased O3 uptake.
17
18 SYMPTOMS OF DISEASE OR TRAUMA, SIGNS OF DISEASE
19 Although insects and diseases are dynamic components of forest ecosystems, trees can be
20 especially susceptible to outbreaks due to the presence of multiple stressors such as drought and
21 pollutant exposure. Ozone can have direct effects on insect or disease organisms, indirect effects
22 on the insect or pathogen through changes to the host, and direct or indirect effects on natural
23 enemies of the insect or pathogen (Pronos et al., 1999). A full discussion of O3 effects on insect
24 and pathogen interactions can be found in Section 9.4.
25 Although the multitude of interacting factors makes it difficult to identify causative factors
26 in the field, some recent examples suggest a role for O3 in the timing or magnitude of disease
27 attacks in the field. After periods of drought stress (such as 1995 in central Europe), the
28 incidence of bark beetle (Ips spp.) appears to increase. In 1998 and 1999, the mean daily capture
29 of Ips was lowest in plots with low O3 exposure; the converse was also true (Grodzki et al.,
30 2002). Elevation confounded the relationships, but the differences in Ips frequency in relation to
31 O3 concentrations were highly significant at lower elevations. In the Valley of Mexico, a 1982
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1 to 1983 drough was documented, but not described as precipitating a bark beetle attack in the
2 early 1980's. However, a link between air pollutant exposure and bark beetle attacks were
3 implicated, because attacked trees were already O3-stressed at the time of the bark beetle attack
4 (Alvarado-Rosales and Hernandez-Tejeda, 2002).
5 An early study showed that oxidant exposure predisposedponderosapine to the root
6 pathogen Fames annosus (James et al., 1980). Both root diseases (Pronos et al., 1999) and O3
7 exposure (Grulke and Balduman, 1999) can each reduce root biomass, leading to increased
8 drought stress, insect attack, and subsequent windthrow or death. Trees may take several years
9 to die, and the patterns of precipitation and annual total rainfall interact to drive the level of
10 drought stress experienced by the tree (Pronos et al., 1999). Additional research is necessary to
11 fully understand the complex interactions occurring between O3 stress and other biotic stresses.
12
13 9.7.5 Ecosystem, Chemical, and Physical Characteristics (water, soil)
14 9.7.5.1 Nutrient Concentrations, Trace Inorganic and Organic Chemicals
15 Ozone exposure reduces the nutritional content of tissues, as well as causing elemental
16 imbalances. Foliar nutrient content may be too high (toxic) or too low (deficient), but the
17 relative amounts and ratios among all nutrients can also result in imbalances.
18 Although N deposition and foliar N content increased with O3 exposure in the San
19 Bernardino Mountains, K, Mg, Fe, and Al were all also higher in ponderosa pine at sites more
20 exposed to air pollution (Poth and Fenn, 1998). Trees with greater foliar injury (due to O3
21 exposure) had higher current year needle concentrations of P, K, Zn, and Fe than trees at the
22 same site that were less injured. In drought-stressed ponderosa pine with O3 exposure, foliar N
23 was also elevated and retained in the remaining needles (Temple et al., 1992). At a relatively
24 clean site in the eastern San Bernardino Mountains, N, P, and K were efficiently readsorbed, but
25 P remaining in the foliage was relatively high compared to defined thresholds. The fact that
26 other elements were modified besides the N being deposited emphasizes the degree of chemical
27 imbalance in the tissue. Foliar micronutrients were within the normal ranges reported for
28 ponderosa pine (Poth and Fenn, 1998; Powers, 1981). Because both N deposition and O3
29 exposure reduce root biomass, it was unlikely that the foliar nutrient content was higher due to
30 greater uptake. Instead, it appaears that retranslocation from senescing tissue was responsible.
31 Across a pollution gradient in the Carpathian Mountains in eastern Europe, only S/N (expected
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1 due to high S deposition) and Fe/Mn ratios were out of balance relative to established norms.
2 The S was relatively high due to SO2 deposition, and the Fe was relatively high due to smelter
3 plumes. No imbalances could be directly attributed to O3 exposure (Mankovska et al., 2004).
4
5 9.7.6 Ecological Processes
6 9.7.6.1 Energy Flow
7 All green plants generate and use energy-containing C compounds through the processes of
8 photosynthesis and respiration. Whole-plant C uptake is dependent on photosynthesis rates, leaf
9 area, and leaf phenology. The effects of O3 at the site of action in the leaf are discussed in
10 Section 9.3. Here, the main focus is on whole-plant carbon dynamics resulting from changes in
11 C acquisition or use under O3 stress.
12 In natural ecosystems, O3 has been shown to depress photosynthesis in sensitive tree
13 species including Pinusponderosa (Miller et al., 1969; Weber et al., 1993; Takemoto et al.,
14 1997; Grulke et al., 2002b) andPopulus tremuloides (Coleman et al., 1995a; Yun and Laurence,
15 1999; Noormets et al., 2001a,b; Sharma et al., 2003). In a study of mature Jeffrey pine, trees in
16 mesic microsites had greater O3 uptake over the growing season in comparison to trees in xeric
17 microsites (Grulke et al., 2003a) and greater O3 uptake was correlated with lower mid canopy
18 needle retention, lower branch diameters, and lower foliar N content (Grulke et al., 2003a).
19 Chamber studies have also shown negative effects of O3 on tree seedling canopy structure
20 (Dickson et al., 2001) and leaf area (Neufeld et al., 1995; Wiltshire et al., 1994). It is well
21 known from chamber and field studies that O3 exposure is correlated with lower foliar retention
22 (Miller et al., 1963; 1972; Karnosky et al., 1996; Grulke and Lee, 1997; Pell et al., 1999; Topa
23 etal., 2001).
24 In contrast to the relatively consistent findings for photosynthesis, O3 effects on respiration
25 have been more variable. Stem respiration was unaffected by O3 exposure (Matyssek et al.,
26 2002), suggesting that construction costs of new stems are not affected by O3. However, the bole
27 represents a relatively large storage pool of carbohydrates in mature trees, and the timing of
28 phenological events among individual trees may help to confound the ability to statistically
29 detect differences in stem respiration across pollutant gradients (Grulke et al., 2001). Below-
30 ground respiration has been found to both increase and decrease in response to O3, depending on
31 the approach and timing of CO2 measures (Coleman et al., 1996; Andersen and Scagel, 1997;
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1 Scagel and Andersen, 1997; King et al., 2001). The decreased soil respiration is thought to be
2 due to reduced root growth under O3 exposure, but could also be partially explained by
3 decreased microbial respiration in response to O3. Additional research is necessary to identify
4 the role of O3 in affecting root vs. heterotrophic respiration, particularly over long time intervals.
5 Carbohydrate availability and use influence the degree to which plants respond to O3
6 exposure. A model simulation of the effect of O3 exposure on bole growth of Pinusponderosa
1 showed a 15% reduction in mass (Weber and Grulke, 1995), largely influenced by differences in
8 carbohydrates allocated and partitioned in repair processes elsewhere in the tree. Foliar
9 respiration is thought to increase under elevated O3 as maintenance costs (energy needs) of
10 leaves damaged by O3 are higher than normal (Grulke and Balduman, 1999; Noormets et al.,
11 200Ib), but differences in foliar respiration are subtle and difficult to detect statistically. Foliar
12 carbohydrate studies also suggest that more C is used under O3 stress for repair processes (Topa
13 et al., 2001; Grulke et al., 2001) which would result in increased respiration. Ozone exposure
14 also reduced enzymatic activities of carbohydrate metabolism related to the breakdown of
15 sucrose (Einig et al., 1997). Changes in soil respiration in response to O3, even though O3 does
16 not penetrate into the soil, illustrates the tight coupling of plant C balance and soil biota and
17 illustrates the potential role O3 plays in altering ecosystem C balances (Andersen, 2000).
18 Ozone can affect plant allometry through changes in energy use, potentially affecting net
19 primary production (NPP) at larger scales. The net effect of O3 impacts on photosynthesis and
20 respiration for sensitive components of natural ecosystems is that height growth (Isebrands et al.,
21 2001; Oksanen, 2003b); and radial growth (Peterson et al., 1987, 1991; Isebrands et al., 2001;
22 Terrazas and Bernal-Salazar, 2002; Oksanen, 2003b) can be negatively affected by O3. This has
23 been extrapolated to decreased NPP (Hogsett et al., 1997; Laurence et al., 2000).
24 Energy flow in plant communities can be altered by O3 through changes in C allocation.
25 It is well known that elevated O3 affects C allocation to roots (Coleman et al., 1995b; Andersen
26 et al., 1997; Grulke et al., 1998; Grulke and Balduman, 1999; Grulke et al., 2001) by decreasing
27 or inhibiting phloem loading of carbohydrates (Grantz and Farrar, 1999; Landolt et al., 1997), or
28 of carbohydrate metabolism (Einig et al., 1997). This leads to depressed root growth (Andersen
29 et al., 1991; Coleman et al., 1996; Grulke et al., 1998) and the potential for plant communities to
30 have an increased susceptibility to drought through altered root-shoot balance. Furthermore, it
31 can negatively affect below ground food webs (Scagel and Andersen, 1997; Phillips et al., 2002).
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1 Another energetically costly response to O3 exposure is that the production of defense
2 compounds, such as antioxidants, tend to increase under elevated O3 conditions (Sheng et al.,
3 1997; Tausz et al., 1999b; 2002). Antioxidants help the plant scavenge free radicals before they
4 can cause damage to membranes or cell walls, but they demand C for production such that
5 growth can be adversely affected. In mature Jeffrey pine, stomatal uptake of O3 elicited one
6 complex of antioxidant defenses in mesic microsites, while endogenously generated free radicals
7 in the chloroplast elicited a second complex of antioxidant defenses in xeric microsites (Grulke
8 et al., 2003b).
9
10 9.7.6.2 Material Flow
11 Plants as producers are responsible for using inorganic atmospheric C and reducing it into
12 organic forms used by consumers, thus driving nutrient processes in ecosystems. Ozone has the
13 potential to disrupt material flow through organic C cycling and changes in nutrient cycling,
14 particularly N and P cycling. Although there is indirect evidence that O3 is disrupting C and
15 nutrient cycling at the ecosystem level, there is little direct evidence that O3 alters nutrient
16 processing at ecosystem scales.
17 The greatest annual nutrient and C input to ecosystems is from foliar and root turnover.
18 Excision of plant parts and whole plant mortality are potentially much larger, but syncopated,
19 ecosystem inputs. Ozone exposure alters C cycling in the ecosystem by affecting the
20 within-plant C allocation and partitioning of dominant, O3-sensitive plants, and through chemical
21 composition and rate of decomposition of sloughed plant parts (roots, branches, leaves)
22 (Figure 9-19).
23 In addition to O3-induced changes in the quantity of C and nutrient inputs into ecosystems,
24 O3 also can alter the nutrient quality of inputs. Ozone exposure alters nutrient levels in the
25 foliage (Boerner and Rebbeck, 1995; Lindroth et al., 2001; Fenn and Poth, 1998; Momen et al.,
26 2002) and affects the C:N ratio (Andersen et al., 2001; Lindroth et al., 2001; Grulke and Lee,
27 1997; Grulke et al., 2003b). Concentrations of compounds such as tannins, lignin and phenolics
28 (Kim et al., 1998; Findlay et al., 1996; Baumgarten et al., 2000; Saleem et al., 2001) are also
29 affected by O3 exposure, which in turn alters decomposability (Fenn and Dunn, 1989) and litter
30 buildup in the ecosystem.
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1 There are several possible pathways by which O3 may affect litter quality and, hence, litter
2 decomposition, thus altering nutrient flow in ecosystems. These include altered C quality,
3 altered nutrient quality, and alteration of leaf surface organisms important in decomposition
4 pathways. For example, yellow poplar (Liriodendron tulipiferd) and black cherry (Prunus
5 serotina) litter exposed to O3 showed greater N loss during decomposition than charcoal filtered
6 controls, although mass loss did not vary among O3 treatments (Boerner and Rebbeck, 1995).
7 Subsequent studies showed that although foliar N was not affected by O3 exposure in yellow
8 poplar leaves, foliage decomposed more slowly (Scherzer et al., 1998). Other studies have also
9 shown a change in foliar N concentration in response to O3 treatment, affecting the C/N ratio and
10 possible litter quality (Andersen et al., 2001).
11 In some cases, it appears that N remobilization from foliage into the plant is not complete
12 at the time of foliage abscission in O3-exposed plants (Findlay and Jones, 1990; Stow et al.,
13 1992; Matyssek et al., 1993; Patterson and Rundel, 1995). Greater N content of senesced litter
14 could increase rates of decomposition. When O3-exposed cottonwood (Populus deltoides) leaves
15 abscissed at the same time as control leaves, they decomposed at similar rates; however,
16 prematurely senesced foliage from O3-exposed cottonwood decomposed more slowly than
17 controls despite their higher N content (Findlay and Jones, 1990; Findlay et al., 1991). Higher N
18 in senesced leaves appeared to be related to organic complexes formed by bound phenolics in
19 O3-exposed leaves, that made the litter less palatable to decomposers, thereby slowing
20 decomposition rates (Jones et al., 1994; Findlay et al., 1996). Increased phenolics also have
21 been found in European silver birch (Betulapenduld) exposed to O3 (Saleem et al., 2001).
22 Carbon quality in leaf litter also changes in O3-exposed foliage. Compositional changes in
23 leaf structural characteristics, such as lignin content, would be expected to alter rates of litter
24 decomposition (Fogel and Cromak, 1977; Meentemeyer, 1978; Kim et al., 1998). Blackberry
25 (Rubus cuneifolus) litter exposed to elevated O3 had greater permanganate lignin than control
26 treatments, which was inversely related to mass-loss rates in decomposition studies (Kim et al.,
27 1998).
28 Ozone may affect early stages of decomposition by altering populations of leaf surface
29 organisms before or after senescence. Magan et al. (1995) found a shift in phyllosphere fungi on
30 Scots pine (Pinus sylvestris\ Sitka spruce (Picea sitchensis\ and Norway spruce (Picea abies)
31 exposed to O3, but the potential effect of these changes on subsequent litter decomposition was
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1 uncertain. The slowest decomposition rates of pre-exposed blackberry leaves were found when
2 senesced foliage was exposed to O3 during decomposition, suggesting a possible direct effect of
3 O3 on microorganisms in decomposing litter (Kim et al., 1998). Whether O3 concentrations at
4 the soil surface influence initial stages of litter decomposition remains to be addressed.
5 Ozone exposure also reduces nutritional content of foliage because of the degradation of
6 chlorophyll. Reconstruction of chlorophyll may be limited by nutritionally poor soils or low soil
7 moisture, as well as alteration of root uptake by O3 exposure and other stressors. Foliar exposure
8 to O3 may also increase leaching of nutrients (Kerstiens and Lendzian, 1989). Ozone exposure
9 promotes early senescence of foliage (Miller and Elderman, 1977; Heath and Taylor, 1997), with
10 higher nutrient content than if excised later in the growing season (Poth and Fenn, 1998).
11 Since O3 can slow decomposition through changes in leaf quality and quantity, leaf litter
12 can accumulate (Fenn and Dunn, 1989). The accumulation of soil organic matter from increased
13 leaf litter, even in the absence of acidic deposition, can lower soil pH (Binkley, 1992). Lower
14 soil pH can promote loss of nutrients, particularly cations, from the system, further reducing
15 nutrient availability to the plant. Complex organic compounds in decomposing litter may also
16 tie up nutrients, rendering them less available to plants.
17 Other factors such as nutrient deposition also affect the degree to which O3 effects C and
18 nutrient flow through ecosystems. In high-pollution sites, the effect of N deposition is difficult
19 to separate from O3 exposure, and reviews of the effect of acidic (and N) deposition on
20 ecosystem nutrient dynamics are important to consider (see Binkley, 1992; Fenn et al., 1998;
21 2003). For example, at moderately high pollution sites, foliar content of N is higher than that at
22 lower pollution site, but so are P, Mg, and Fe contents (Poth and Fenn, 1998). Although
23 significant changes in foliar tissue chemistry have occurred in response to long-term pollutant
24 deposition in the Carpathians (Mankovska et al., 2002; Fenn et al., 2002), much of this response
25 is correlated to heavy metal and N and S deposition. At this point, the contribution of O3
26 exposure alone cannot be isolated without careful between-site comparisons of plant response
27 and understanding the deposit!onal velocities of constituent atmospheric species. The significant
28 effects of plant response along known pollution gradients are important to consider because
29 heavy metal and N and S deposition has significantly declined over the last decade, while O3
30 exposures remain high and declines in forest health have been sustained.
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1 Models provide a means to track material flows through ecosystems. Two biogeochemical
2 models were parameterized to capture long-term effects of O3 exposure, N deposition, and
3 climate on aponderosapine-dominated site in the eastern San Bernardino Mountains.
4 Simulated O3 exposure resulted in faster production and turnover of foliage and a shift in carbon
5 from the canopy (15% reduction) to the forest floor (increase of 50 to 60%) (Arbaugh et al.,
6 1999). When O3 exposure was combined with that of N deposition, litter mass exponentially
7 increased.
8 The direct effect of O3 exposure on below-ground nutrient dynamics and ecosystem
9 material flow is poorly understood. Additional research will be necessary to understand spatial
10 and temporal dynamics of nutrient and C flow in ecosystems and to separate the effects of O3
11 from those attributable to N deposition.
12
13 9.7.7 Hydrological and Geomorphological
14 At present, there are no publications on the effects of O3 exposure that are carried through
15 at the ecosystem level to changes in mass water flow, channel morphology, riparian habitat
16 complexity, or sediment movement. It is possible that processes occurring at smaller scales are
17 affecting geomorphological processes in ecosystems; however, difficulty in scaling these
18 responses spatially and temporally have made it difficult to show experimentally. It is possible
19 that O3 exposure affects water quality through changes in energy and material flows, as
20 discussed previously.
21
22 9.7.8 Natural Disturbance Regimes
23 There has been little research on how natural disturbances interact with O3 to affect
24 performance of plants, communities, and ecosystems. The frequency, intensity, extent, and
25 duration of natural disturbances are variable and unpredictable. However, there have been
26 enough ecophysiological studies to suggest that O3 could predispose plant communities to
27 certain natural stresses, e.g., drought, stress, or extreme low temperature stress during the winter.
28 While several studies have shown that drought stress reduces O3 uptake through stomatal
29 closure, evidence also suggests that O3 can alter plant water use and susceptibility to drought.
30 In controlled studies, Reich and Lassoie (1984) showed that relatively low O3 concentrations
31 could diminish stomatal control and alter water use efficiency. Ash trees (Fraxinus excelsior}
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1 exposed to elevated O3 had greater water use early in the growing season, but less water use late
2 in the growing season when exposed to elevated O3 (Wiltshire et al., 1994). Under moderate
3 drought stress, Picea abies trees grown under elevated O3 consumed water faster and showed
4 higher stomatal conductances than controls (Karlsson et al., 1995). Pearson and Mansfield
5 (1993) showed that successive O3 episodes disrupted stomatal function, making Fagus sylvatica
6 seedlings more susceptible to drought. Previous year O3 exposure was shown to have a
7 carryover effect in the following growing season for Fagus sylvatica (Pearson and Mansfield,
8 1994).
9 Few studies showing the effects of O3 on water relations of field-grown trees are found in
10 the literature. However, Grulke et al. (2003a) examined the effects of O3 on canopy transpiration
11 of Pinus Jeffreyi from mesic and xeric microsites and found that trees from mesic sites had 20%
12 more O3 uptake than those in the xeric sites. The authors also concluded that the mesic trees had
13 greater O3 injury as evidenced by lower needle retention, whereas trees in xeric microsites had
14 greater chlorotic mottle. Chlorotic mottle induced by stomatal uptake of O3 is indistinguishable
15 from that of endogenously produced oxidants resulting from partially closed stomata, a reduction
16 of CO2 inside the leaf, and production of strong oxidizers within the chloroplast when excited
17 electrons are passed to O2 instead of CO2 under high light levels.
18 Trees living near the limits of their freezing-tolerance range may be especially susceptible
19 to predisposition of freezing injury by O3 (Sheppard et al., 1989). However, Pinus halepensis
20 exposed to elevated O3 had enhanced winter hardiness (Wellburn and Wellburn, 1994). As with
21 the seasonal carryover of drought susceptibility, the influence of elevated O3 on freezing
22 tolerance is carried over from summer to winter. Such effects have been demonstrated for Picea
23 sitchensis (Lucas et al., 1988) and for Picea rubens (Waite et al., 1994). Sorting out the role of
24 elevated O3 in contributing to frost or low-temperature damage in forests remains difficult due to
25 the presence of other factors that may affect senescence.
26
27 9.7.9 Scaling to Ecosystem Levels
28 The vast majority of literature describing O3 effects comes from short-duration herbaceous
29 plant or tree seedling studies under controlled conditions. Scaling results from these studies
30 requires extrapolation over both space and time in order to understand the full extent of changes
31 in ecosystems. In addition to spatial, temporal, and age-related complexities, ecosystems are
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1 composed of organisms whose lifetimes range from hours to centuries (Laurence and Andersen,
2 2003). Forested ecosystems are affected by environmental conditions such as water and nutrient
3 availability, as well as by intra- and interspecies competition. Therefore, direct experimentation
4 to determine the response of forested ecosystems is not simply a matter of determining the effect
5 of O3 on individual mature trees. In addition, even if an experiment can be conducted,
6 extrapolation of the results across landscapes and regions remains challenging. Nonetheless,
7 models provide a means to explore possible long-term changes and to identify important
8 research uncertainties.
9 Approaches to scaling fall roughly into two categories: (1) process-based modeling to
10 extrapolate physiological responses to O3 based on seedling studies and (2) field assessments
11 using surveys and growth correlations, often in association with stand-level models to address
12 ecosystem complexity. Comparatively good information is available on process level effects of
13 O3 in seedlings and, therefore, some models offer the opportunity to use this information to scale
14 O3 effects at the stand and regional scales (Hogsett et al., 1997; Fuhrer et al., 1997; Chappelka
15 and Samuelson, 1998; Laurence et al., 2000).
16
17 9.7.9.1 Scaling from Seedlings to Mature Trees
18 A number of investigators have used simulation models based on physiological processes
19 to integrate available data and predict the effects of O3 on mature trees. Such models predict tree
20 growth by simulating fundamental mechanisms, rather than through a statistical analysis of
21 empirical data. For instance, the process of photosynthesis is simulated based on environmental
22 conditions and physiological characteristics, and then the fixed C is allocated to plant growth
23 using principles of plant physiology. Models based on mechanisms should be applicable across
24 wide areas if the important functional relationships are represented accurately in the models and
25 if the environmental conditions are accurately identified. The ability of six models (TREGRO,
26 CARBON, ECOPHYS, PGSM, TREE-BGC, and W91) to simulate the effects of climatic
27 change and O3 have been reviewed by Constable and Friend (2000). Of these models, only
28 PGSM and TREGRO explicitly simulated the effects of O3 on foliar processes.
29 The TREGRO model was used to simulate C allocation and tissue growth in seedlings and
30 mature red oak trees based on the experimental data discussed above (Weinstein et al., 1998).
31 For seedlings at 2x-ambient O3, only the total nonstructural carbohydrate (TNC) storage pool
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1 was predicted to be affected. For mature trees, large decreases were predicted for TNC, leaves,
2 stem, branch, and both fine and coarse roots. Most predicted effects in mature trees were
3 consistent with observations in the field, but the simulations overestimated the effect of
4 2x-ambient O3 on root TNC and growth. The authors suggested that this discrepancy may
5 have been due to trees reducing respiration in response to O3 stress, a response not simulated in
6 the model.
7 For Abies concolor, TREGRO was parameterized and simulated growth of a mature tree
8 for 3 years to test for effects of O3 exposure and drought stress (Retzlaff et al., 2000).
9 Reductions in O3 exposure-mediated carbon assimilation were translated to losses in whole tree
10 biomass that probably would not be detectable in the field. However, TNC levels in branch
11 tissue were simulated to be lowered by over 50%, and branch growth was reduced in a
12 moderately polluted site relative to a clean site. Low O3 exposure (sufficient to decrease C
13 assimilation by 2.5%) and drought stress (25% reduction in annual precipitation, which is
14 common on a decadal scale) acted synergistically to reduce C gain of whole tree biomass of
15 A. concolor. Simulated results of the tests were comparable to effects found in OTCs for
16 seedlings and pole-sized trees in clean and moderately polluted sites.
17 Models such as TREGRO are usually parameterized from many different sources of data,
18 including chamber experiments and plantations, from seedlings to mature trees, making it
19 difficult to validate that they reproduce changes that occur as trees develop from seedlings to
20 maturity. To address this issue, physiological and growth data were collected from a natural
21 stand of P. ponderosa and used to parameterize the TREGRO model (Grulke and Retzlaff,
22 2001). Representative trees of each of five tree age classes were selected based on population
23 means of morphological, physiological, and nearest neighbor attributes. Seedlings were
24 observed to differ significantly from pole-sized and older trees in most physiological traits. The
25 changes in biomass with tree age predicted from the model closely matched those of trees in the
26 natural stand.
27 The PGSM model was used to simulate Pinus ponder osa seedling growth responses to O3
28 exposure and drought stress (Chen et al., 1994). Drought stress was predicted to reduce the
29 effect of O3 on growth, as was observed in the experimental data. The TREGRO model was
30 used to simulate responses of Pacific Coast and interior varieties of P. ponder osa to five
31 simulated O3 exposures between subambient and 3 x-ambient (Constable and Taylor, 1997).
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1 Simulated growth of var. ponderosa was reduced more than that in var. scopulorum with all O3
2 exposures. Drought was protective of O3 exposure. Similar results were also found in a
3 relatively moist ponderosa pine plantation (Panek and Goldstein, 2001), whereas drought was
4 synergistically deleterious with cumulative O3 exposure in a natural stand (Grulke et al., 2002b).
5 For Pinus taeda, the Plant Growth Stress Model (PGSM) was calibrated with seedling data
6 and then used to simulate the growth of mature trees over a 55 year period in the Duke Forest,
7 NC, using estimates of historical O3 concentrations (Chen et al., 1998). Simulated stem diameter
8 and tree height were comparable to observed values. In another simulation using TREGRO,
9 loblolly pine was more sensitive (greatest reduction in C gain) to a peak O3 episode in July
10 (Constable and Retzlaff 1997), whereas mature tulip poplar (Liriodendron tulipiferd) was more
11 sensitive to a peak O3 episode in August.
12 For Populus tremuloides, the ECOPHYS model was used to simulate the relative above-
13 ground growth response of an O3-sensitive clone (259) exposed to square-wave variation in O3
14 concentration (Martin et al., 2001). The model adequately simulated several effects of O3,
15 including a greater effect on stem diameter than on stem height, earlier leaf abscission, and
16 reduced stem and leaf dry matter production at the end of the growing season. For Acer
17 sacchamm, the TREGRO model was use to predict effects of a 10-year O3 exposure on root and
18 stem growth of a simulated 160-year-old tree (Retzlaff et al., 1996). Twice-ambient O3 exposure
19 (for Ithaca, NY) was predicted to deplete the TNC pools and reduce fine root production.
20
21 9.7.9.2 Surveys, Growth Correlations and Stand-Level Modeling
22 Stand-level studies have included surveys of O3 symptoms, correlations of radial growth
23 with O3 and other environmental factors, and regional scale modeling. In addition, open air O3
24 exposure systems, such as those being used on Populus tremuloides-mixed stands in northern
25 Wisconsin (Karnosky et al., 2003a) and on Fagus sylvatica and Picea abies in Germany (Nunn
26 et al., 2002) offer an opportunity to examine larger plot sizes, older trees, and trees growing
27 under realistic competition. Plots along natural O3 gradients, as have been used very effectively
28 in southern California forest studies (Miller and McBride, 1999) and with P. tremuloides stands
29 in the Great Lakes region (Karnosky et al., 1999), offer additional insights into ecosystem level
30 responses. Undoubtedly, however, simulation modeling will have to become an integral
31 component of research in order to predict adequately ecosystem responses to O3 (Laurence and
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1 Andersen, 2003). Results of these approaches are discussed below, organized into three U.S.
2 regions: (1) northern states (including the upper Midwest and the Northeast), (2) southeastern
3 states, and (3) western states (primarily California). A fourth section contains selected
4 information from Europe.
5
6 Northern and Midwestern United States. In recent years, the USD A Forest Service has
7 conducted systematic O3 biomonitoring surveys in most north-central and northeastern states
8 (Smith et al., 2003; Coulston et al., 2003). Plots are located on a systematic grid, and trained
9 field crews evaluate up to 30 plants of up to six species that have foliar injury symptoms
10 diagnostic of O3 damage. For the United States as a whole, injury has been found more often in
11 eastern than in interior or west-coast states. As expected, O3 injury is more common and more
12 severe in areas with higher O3 concentrations. Of sampled Prunus serotina plots, -12% were
13 estimated to be at high risk for injury based on a injury index derived from the survey data
14 (Coulston et al., 2003). P. serotina was estimated to be at risk for injury on the Allegheny
15 Plateau and the Allegheny Mountains (in Pennsylvania, West Virginia, and Maryland), as well as
16 in the coastal plain of Maryland and Virginia.
17 Ozone concentration, foliar injury, and physiological traits were measured on P. serotina
18 trees of different sizes in Pennsylvania (Fredericksen et al., 1995). The proportion of foliage
19 injured was 46% for seedlings, 15% for saplings, and 20% for canopy trees. Cumulative O3 flux
20 was the most useful O3 metric for predicting injury. Injury was negatively correlated (r2 = 0.82)
21 with net photosynthetic rates, but was not related to stomatal conductance. P. serotina is
22 discussed further below, because detailed surveys have been conducted in the Shenandoah and
23 Great Smoky Mountains National Parks in the southeastern United States.
24 Over the past several decades, some surveys ofPinus strobus have reported significant
25 associations between foliar injury and reduced growth (Anderson et al., 1988; Benoit et al.,
26 1982). However, a review of 93 surveys conducted from 1900 through the late 1980s concluded
27 that methodological problems were pervasive, including such issues as proximity to roads, lack
28 of peer review, lack of random sampling, small sample sizes, and lack of quantitative methods to
29 estimate severity (Bennett et al., 1994). Because of these problems, along with evidence of
30 adequate growth rates for P. strobus regionally and contradictory evidence from numerous
31 studies of symptom production in response to controlled O3 exposure, these authors concluded
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1 that there was no clear evidence of decline in P. strobus (Bennett et al., 1994). A more recent
2 study in Acadia National Park (ANP) in Maine found no association between O3 exposure in
3 OTCs and symptom development in P. strobus, calling into question whether symptoms
4 previously ascribed to O3 may be caused by some other stress (Kohut et al., 2000). However,
5 another ANP study found significant correlations between O3 exposure and the radial growth of
6 trees during 10 years in 7 of 8 stands examined (102 trees total; Bartholomay et al., 1997).
7 Taken together, these results suggest that there may not an association between growth of P.
8 strobus trees and putative O3 symptoms, but there may be an association between O3 exposure
9 and radial growth of mature trees in the field.
10 The study of O3 effects was undertaken from 1990 to 1993 in the ANP in Maine, because
11 this location experiences elevated O3 exposures due to transport from urban areas located upwind
12 (Kohut et al., 2000). Thirty-two species of plants found in the park were propagated and
13 exposed to O3 in OTCs. In addition, ambient O3 concentrations were monitored at the study site
14 at 15 m above sea level and near the top of Cadillac Mountain at 470 m above sea level. At the
15 study site, the maximum 1-h O3 concentration was 140 ppb, which occurred in both 1990 and
16 1991. Daytime 12-h O3 concentrations were 35, 41, 36, and 37 ppb during the four years; and O3
17 concentrations were consistently higher at the high-elevation site. Species showing foliar injury
18 at ambient O3 concentrations included Prunus serotina, Populus tremuloides, Fraxinus
19 americana, Pinus banksiana, big-leaf aster, and spreading dogbane. Species showing foliar
20 injury at 1.5 x ambient O3 concentrations included Betula populifolia, small sundrops, and
21 bunchberry. Species remaining uninjured at 2 x ambient O3 concentrations included Betula
22 papyri/era., Pinus strobus, Pinus rigida, Picea rubens, Thuja occidentalis, Quercus robur,
23 Canada bluejoint grass, wild radish, and Canada mayflower. Because of their O3 sensitivity and
24 diagnostic symptoms, big-leaf aster, spreading dogbane, Populus tremuloides, Fraxinus
25 americana, and Prunus serotina were recommended as bioindicators for the ANP.
26 The PnET-II model was applied to 64 locations across the northeastern United States to
27 simulate the effects of ambient O3 on mature hardwood forests (Ollinger et al., 1997). In this
28 model, O3 effects on each of several layers of the forest canopy were represented by a single
29 linear equation relating predicted O3 uptake to decreased net photosynthetic rate. Wood growth
30 was predicted to decrease between 3 to 22%, with greatest reductions in southern portions of the
31 region where O3 levels were highest and on soils with high water-holding capacity where
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1 drought stress was absent. Little variation was predicted among years, because high O3 often
2 coincided with hot, dry weather conditions that reduced predicted stomatal conductance and
3 O3 uptake.
4 In order to estimate the impact of O3 on forests, effects must be evaluated not only on
5 individuals, but also on mixtures of species and the composition of forest stands. The PnET
6 model described above evaluated the effects of O3 on broad forest types (an evergreen/deciduous
7 mix), but did not address specific forest species composition. In order to address competition
8 among species, the TREGRO model was linked to the ZELIG forest stand growth model and a
9 geographic information system was created to predict the effects of O3 across the north-central
10 and northeastern United States (Laurence et al., 2000). ZELIG is a gap-succession model used
11 to simulate succession in mixed stands typical of eastern and northern forests. Ambient O3
12 generally caused a reduction of 2 to 4% in the growth ofQuercus robur across the region during
13 the 100-year simulation. The response followed the pattern of O3 exposure, with little effect in
14 the northwest part of the region, but with greater effect in southern locations. The O3 response of
15 Acer saccharum to O3 varied widely, but the overall growth response was always positive,
16 indicating that the evergreen/deciduous mix was able to take advantage of the decrease in the
17 growth of Q. robur and other species caused by O3. In the northernmost part of the region,
18 A. saccharum growth increased by up to 3%, but in the southern part of the region, its growth
19 increased by up to 12%. The authors ascribed this enhanced growth to a combination of warmer
20 temperatures and reduction in the growth of Primus serotina, a minor component of the
21 simulated stand that was very sensitive to O3.
22
23 Southeast United States. In a survey of the Great Smoky Mountains National Park, foliar
24 injury attributed to O3 was found on 47% of the more than 1,600 plants examined (Chappelka
25 et al., 1997). In subsequent surveys of injury in the park, injury was found on mature trees of the
26 following species: Sassafras albidum, Prunus serotina, and Liriodendron tulip/era (Chappelka
27 et al., 1999a,b). In a similar study in Shenandoah National Park, injury was found on Fraxinus
28 americana, P. serotina, andZ. tulip/era (Hildebrand et al., 1996).
29 For Prunus serotina seedlings grown in soil in OTCs and exposed to relatively low
30 ambient levels of O3 in Pennsylvania, there was no correspondence between visible foliar stipple,
31 leaf gas exchange, and seedling growth between two families previously shown to differ in O3
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1 symptoms (Kouterick et al., 2000). However, significant exposure-response relationships were
2 found for foliar injury in the Great Smoky Mountains and Shenandoah National Parks. In each
3 park, foliar injury was evaluated on mature P. serotma trees on three plots at different elevations
4 near O3 monitors during 1991 to 1993 (Hildebrand et al., 1996; Chappelka et al., 1999a,b).
5 In 1991, incidence was 60% and 45% for the two parks and 33% in both parks during 1992 and
6 1993. Symptoms were greater at the highest elevations where O3 concentrations were highest.
7 In another study, radial growth rates were measured in 44 P. serotma trees ranging in age from
8 19 to 56 years old with and without O3 symptoms, at three sites in the Great Smoky Mountains
9 National Park. Trees with O3 symptoms were compared to similar-sized trees with few
10 symptoms. There was no evidence that trees with O3 symptoms had lower growth rates (p = 0.6)
11 (Somers et al., 1998).
12 In the Great Smoky Mountains National Park, radial growth rates were measured for
13 44 L. tulip/era trees ranging in age from 30 to 58 years old, with and without O3 symptoms, at
14 three sites at different elevations (Somers et al., 1998). Trees with O3 symptoms averaged 30%
15 lower growth rates over ten years (p = 0.0005). Seedlings of Liriodendron tulip/era were
16 exposed for two seasons to 2x-ambient O3 exposures in OTCs in Delaware, OH (seasonal
17 SUMOO exposures of 107 and 197 ppmh) (Rebbeck, 1996). Foliar O3 symptoms were observed,
18 but growth was not reduced.
19 In order to evaluate the influence of interspecies competition on O3 effects, the linked
20 TREGRO and ZELIG modeling system was used to predict the effects of O3 over 100 years on
21 the basal area of species in a Liriodendron tulipfera-dominated forest in the Great Smoky
22 Mountains National Park (Weinstein et al., 2001). Ambient O3 was predicted to reduce the basal
23 area of L. tulip/era by 10%, whereas a 1.5x-ambient exposure was predicted to cause a 30%
24 reduction. Basal area of Acer rubrum and Prunus serotma was predicted to increase for some
25 years, but then decrease by up to 30%, with few changes in the total basal area of all species by
26 the end of the simulation.
27 In order to evaluate the influence of interspecies competition on O3 effects, the linked
28 TREGRO and ZELIG modeling system was used to predict the effects of O3 on the basal area of
29 Pinus taeda and Liriodendron tulip/era growth throughout their ranges (Laurence et al., 2003).
30 The models were parameterized using biological and meteorological data from three sites in the
31 southeastern United States (in Alabama, Louisiana, and North Carolina). Forest stand response
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1 to five O3-exposure regimes with annual SUM06 values ranging from 0 to 100 ppmh per year
2 was simulated for 100 years. The simulated basal area of the two species was generated within
3 the context of four other tree species common in southeastern forests. Basal area of P. taeda was
4 highly responsive to precipitation and O3 exposure, with the greatest increases under high-
5 precipitation, lowO3-exposure scenarios and the greatest decreases under low-precipitation, high
6 O3-exposure scenarios. The basal area of L. tulipifera did not significantly differ (+10%) from
7 simulations using a "base case" (ambient O3, average precipitation).
8 Systematic biomonitoring surveys found that approximately 24% of sampled sweetgum
9 and 15% of sampled Pinus taeda plots were estimated to be at high risk for foliar injury on the
10 coastal plain of Maryland and Virginia (Smith et al., 2003; Coulston et al., 2003). In a study in
11 Tennessee, the effect of ambient (uncontrolled) O3 on 28 mature canopy-dominant 50 to
12 90-year-old P. taeda trees in five stands was measured over a 5-year period (McLaughlin and
13 Downing, 1995, 1996). Of many O3 metrics, a 3-day average of hourly O3 values > 40 ppb
14 (AOT40) was found to best explain short-term variation in stem expansion as measured with
15 dendrometer bands. Interactions between O3, temperature, and drought stress (as indicated by
16 the weekly moisture stress index) accounted for 63% of the short-term variation in stem growth
17 rates. Because there are interactions among O3, drought stress, and temperature that may differ
18 with the averaging time (days to years), this type of study cannot provide conclusive proof of
19 cause and effect (Reams et al., 1995). However, the results do suggest that the effects of O3
20 measured on loblolly seedlings may also be occurring in mature trees in both wet and dry sites.
21 The magnitude of effects of O3 on growth, including interactions with other variables in this
22 study, ranged from 0 to 15% over 5 years, with an average of 5.5%.
23
24 Western United States. The USD A Forest Service conducted O3 biomonitoring surveys in
25 Washington, Oregon, and California during one year (1998), and in the Sierra Nevada and
26 Sequoia National Forest every other year for several decades (Campbell et al., 2000). Overall,
27 only one plot showed any symptoms of O3 injury outside of the Sierra Nevada and Sequoia
28 National Forests. In the Sierra Nevada National Forest, between 30 and 40% of trees showed
29 injury from 1989 through 1997. In the Sequoia National Forest, between 40 and 50% of the
30 trees surveyed showed injury from 1990 through 1998.
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1 For Pinus ponderosa along a well-studied gradient of O3 exposure in the San Bernardino
2 mountains, chlorotic mottle was highest on foliage at the most polluted site, as has been found
3 previously (Grulke and Balduman, 1999). Based on whole-tree harvests, root biomass was
4 lowest at the most polluted sites, confirming previous studies with seedlings under controlled
5 conditions, as discussed above. Ozone responses in highly polluted environments such as
6 Southern California may not be predicted adequately by extrapolating effects from single-factor
7 experiments. Instead, combined approaches utilizing field experiments and modeling efforts
8 may be required to properly account for a combination of stressors including O3, N deposition,
9 and drought. Furthermore, the available studies underscore the lack of correlation between O3
10 symptoms and mature tree effects. If a field survey fails to find a correlation between mature
11 tree growth and O3, this result may be due to the dominant effect of another factor such as N
12 deposition and may not be evidence that O3 does not reduce the growth of mature trees.
13 The TREGRO model was used to evaluate how projected future temperature and CO2
14 concentrations might affect the response of individual ponderosa pine to O3 at seven sites in
15 California, Oregon, and Washington (Tingey et al., 2001). As expected, growth decreased with
16 increasing O3 exposure. Differences in O3 response among sites appeared to be due primarily to
17 differences in precipitation.
18 Often air quality standards do not translate directly into measureable improvements in tree
19 growth or productivity. To evaluate whether past improvements in air quality have improved
20 ponderosa pine growth, TREGRO was used to simulate growth at sites in the San Bernardino
21 Mountains in California (Tingey et al., 2004). Ozone and meteorogical data from the past
22 37 years was used to run the simulations. Despite variation in precipitation and temperature,
23 O3 was found to reduce simulated tree growth. The authors were able to simulate growth
24 improvements as air quality improved during the 1980s and 1990s, suggesting that
25 improvements in emission control strategies benefited ponderosa pine. The model simulations
26 were qualitatively consistent with improvements in canopy condition that were observed at sites
27 where O3 reductions were the greatest.
28
29 Studies in Europe. In a 4-year study ofFagus sylvatica in Switzerland at 57 forest sites
30 ranging in age from 65 to 173 years, stem increment was found to decrease with increasing
31 maximum O3 exposure (Braun et al., 1999). In this study, O3 concentration was estimated by
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1 interpolation among monitoring stations, and other site conditions such as soil water status and
2 temperature were interpolated from weather stations. Other factors such as N deposition, tree
3 diameter, and canopy dominance were also found to be significantly associated with stem
4 increment. The maximum annual O3 dose (expressed as AOT40) was found to be more strongly
5 associated with decreased stem increment than was the average O3 dose over the 4 years.
6 A growth reduction of 22.5% (confidence interval 14.3 to 28.6%) was associated with each
7 10 ppmh increment of O3 (expressed as AOT40). This decrease was steeper than the 6.1%
8 growth reduction summarized previously from several OTC studies with F. sylvatica seedlings
9 (Fuhrer et al., 1997). However, the authors suggest that this difference may be explained largely
10 by the 4 years of exposure in their forest survey study as compared to the one-year exposures for
11 seedlings. As with any forest survey, these results must be interpreted with caution because O3
12 exposure was correlated with other variables, such as tree age and the deposition of NO2
13 and SO2.
14
15 9.7.10 Summary of Ecological Effects of Ozone Exposure on
16 Natural Ecosystems
17 In this chapter, an effort has been made to discuss the adverse effects of O3 on natural
18 ecosystems within the context of the SAB framework for assessing and reporting ecological
19 conditions (Young and Sanzone, 2002). Using this framework, there is evidence that
20 tropospheric O3 is an important stressor of natural ecosystems, with well-documented impacts on
21 the biotic condition, ecological processes, and chemical/physical nature of natural ecosystems
22 (Figure 9-20). In turn, the effects of O3 on individual plants and processes are scaled up through
23 the ecosystem affecting processes such as energy and material flow, intra and interspecies
24 competition, and NPP. Thus, effects on individual keystone species and their associated
25 microflora and fauna may cascade through the ecosystem to the landscape level. This suggests
26 that by affecting water balance, cold hardiness, tolerance to wind, and by predisposing plants to
27 insect and disease pests, O3 may even influence the occurrence and impact of natural
28 disturbance. Despite the probable occurrence of such effects, however, there are essentially no
29 instances where highly integrated ecosystem-level studies have conclusively shown that O3 is
30 indeed altering ecosystem structure and/or function.
31
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Hydrolie Alteration
Habitat Conversion
Habitat Fragmentation
Climate Change
Invasive Non-native Species
Turbidity/Sedimentation
Pesticides
Disease/Pest Outbreaks
Nutrient Pulses
Metals
Dissolved Oxygen Depletion
Ozone (Tropospheric)
Nitrogen Oxides
Nitrates
Hydroiic Alteration
Habitat Conversion
Habitat Fragmentation
Climate Change
Over-Harvesting Vegitation
Large-Scale Invasive
Species Introduction
Large-Scale Disease/Pest
Outbreaks
Hydroiic Alteration
Habitat Conversion
Climate Change
Over-Harvesting Vegitation
Disease/Pest Outbreaks
Altered Fire Regime
Altered Flood Regime
Hydroiic Alteration
Habitat Conversion
Climate Change
Turbidity/Sedimentation
Pesticides
Nutrient Pulses
Metals
Dissolved Oxygen Depletion
Ozone (Tropospheric)
Nitrogen Oxides
Nitrates
Sulfates
Salinity
Acidic Deoosition
t/J
£
+•>
w
Hydroiic Alteration
Habitat Conversion
Climate Change
Pesticides
Disease/Pest Outbreaks
Nutrient Pulses
Dissolved Oxygen Depletion
Nitrogen Oxides
Nitrates
Sylfates
Hydroiic Alteration
Habitat Conversion
Habitat Fragmentation
Climate Change
Turbidity/Sedimentation
Figure 9-20. Common anthropogenic stressors and the essential ecological attributes
they affect.
Source: Modified from Young and Sanzone (2002).
1 Systematic injury surveys demonstrate that foliar injury occurs on sensitive species in
2 many regions of the United States. However, the frequent lack of correspondence between foliar
3 symptoms and growth effects means that other methods must be used to estimate the regional
4 effects of O3 on tree growth rates. Investigations of the radial growth of mature trees in
5 combination with data from many controlled studies with seedlings, as well as a few studies with
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1 mature trees suggest that ambient O3 is reducing the growth of mature trees in some locations.
2 Studies using models based on tree physiology and forest stand dynamics suggest that modest
3 effects of O3 on growth may accumulate over time and may interact with other stresses. For
4 mixed-species stands, such models predict that overall stand growth rate is generally not likely to
5 be affected. However, competitive interactions among species may change as a result of growth
6 reductions of sensitive species. These results suggest that O3 exposure over decades may be
7 altering the species composition of forests in some regions.
8
9 RESEARCH NEEDS
10 The knowledge base for examining the range of ecological effects of O3 on natural
11 ecosystems is growing, but significant uncertainties remain regarding O3 effects at the ecosystem
12 level. For example, there is a need for information on the following ecosystem-level responses:
13 • Ecosystem Processes. Little is known about the effects of O3 on water, C, and nutrient
cycling, particularly at the stand and community levels. Effects on belowground
ecosystem processes in response to O3 exposure alone and in combination with other
stressors are critical to projections at the watershed and landscape scales. Little is yet
known about the effects of O3 on structural or functional components of soil food webs,
or how these impacts could affect plant species diversity (Andersen, 2003).
14 • Biodiversity and Genetic Diversity. The study of genetic aspects of O3 impacts on natural
ecosystems has been largely correlational in nature and it remains to be shown
conclusively whether O3 affects biodiversity or genetic diversity (Pitelka, 1988; Winner
et al., 1991; Davison and Barnes, 1998). Studies of competitive interactions under
elevated O3 levels are needed (Laurence and Andersen, 2003), and reexamination via new
sampling of population studies to bring in a time component to previous studies showing
spatial variability in population responses to O3 are needed. These studies could be
strengthened by modern molecular methods to quantify impacts on diversity.
15 • Natural Ecosystem Interactions with the Atmosphere. Little is known about feedbacks
between O3 and climate change on volatile organic compound (VOC) production, which in
turn, could affect O3 production (Fuentes et al., 2001). At moderate-to-high O3 exposure
sites, aberrations in stomatal behavior could significantly affect individual tree water
balance of sensitive trees, and if the sensitive tree species is dominant, hydrologic balance
at the watershed and landscape level could be affected. This has not been addressed in any
model because O3 exposure effects, if included in the modeling effort have assumed a
linear relationship between assimilation and stomatal conductance.
16 • Below-Ground Inter actions. While the negative effects of O3 on below ground growth are
well characterized, interactions of roots with the soil or microorganisms are not.
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Other Interactive Effects. Interaction studies with other components of global change
(e.g.,warming, increasing atmospheric CO2, N deposition, etc.) or with various biotic
stressors are needed to better predict complex interactions likely in the future (Laurence
and Andersen, 2003). Whether O3 will negate the positive effects of an elevated CO2
environment on plant carbon and water balances is not yet known; nor is it known if these
effects will scale up through the ecosystem. How might O3 affect the progress of pest
epidemics and insect outbreaks as concentrations increase is unclear (Skarby et al., 1998).
Reproduction Effects. Information concerning the impact of O3 on reproductive processes
and reproductive development under realistic field or forest conditions are needed as well
as examination of reproductive effects under interacting pollutants (Black et al., 2000).
Comparative Extrapolation. The vast majority of O3 studies of trees have been conducted
with young, immature trees and in trees that have not yet formed a closed canopy.
Questions remain as to the comparability of O3 effects on juvenile and mature trees
and on trees grown in the open versus those in a closed forest canopy in a competitive
environment (Chappelka and Samuelson, 1998; Kolb and Matyssek, 2001; Samuelson
and Kelly, 2001).
Scaling-Up Issues. Scaling the effects of O3 from the responses of single or a few plants
to effects on communities and ecosystems is a complicated matter that will require a
combination of manipulative experiments with model ecosystems, community and
ecosystem studies along natural O3 gradients, and extensive modeling efforts to project
landscape-level, regional, national and international impacts of O3. Linking these various
studies via impacts on common research quantification across various scales using
measures of such factors as leaf area index or spectral reflective data, which could
eventually be remotely sensed (Kraft et al., 1996; Panek et al., 2003), would provide
powerful new tools for ecologists.
Comparative Risk Assessment Methodologies. Methodologies to determine the important
values of services and benefits derived from natural ecosystems such that these could be
used in comprehensive risk assessment for O3 effects on natural ecosystems (Heck et al.,
1998).
8 9.8 ECONOMIC EVALUATION OF OZONE EFFECTS ON
9 AGRICULTURE, FORESTRY AND NATURAL ECOSYSTEMS
10 9.8.1 Introduction
11 The adverse consequences of ambient of air pollutant exposures on vegetation, ecosystems
12 and components of the material environment have been documented since the beginning of the
13 industrial revolution. Attempts to quantify the monetary damage and injury resulting from
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1 tropospheric O3 exposures to managed agriculture, forests, and natural ecosystems date back at
2 least to the 1950s.
3 Both methodological and data problems plagued early efforts to assess the monetary
4 damages of air pollution to crops and natural vegetation. Adams and Crocker (1989) discussed
5 the methodological issues, e.g., a lack of reliable data on effects from air pollutants on crop
6 yields or the failure to develop and apply appropriate economic models. Some of these problems
7 were remedied by the EPA's National Crop Loss Assessment Network (NCLAN) in the 1980s.
8 The EPAs NCLAN facilitated the performance of economic assessments by providing O3-crop
9 yield data with which to estimate O3-crop yield response functions (see Heagle [1988] for a
10 review of NCLAN procedures and findings). NCLAN also funded a series of economic
11 assessments that, along with subsequent economic assessments, documented substantial
12 economic damages to agriculture. (See Spash [1997] for a detailed review of economic
13 assessments, many of which used NCLAN data.)
14 Since the completion of the NCLAN program in the late 1980s, the number of economic
15 assessments of air pollution studies focusing on terrestrial ecosystems in general, and agriculture
16 in particular, has declined. For example, for the period of 1980 to 1990, 33 economic studies of
17 O3 and other air pollutant effects on U.S. crops were published in peer-reviewed journal outlets
18 (Spash, 1997). However, in preparing this section of the current criteria document, only four
19 peer-reviewed economic assessments were found for the decade of 1991 to 2000 that addressed
20 vegetation in the United States. In addition, one peer-reviewed article (Kuik et al., 2000) was
21 found dealing with agriculture in the Netherlands. Recent interest in global climate change and
22 the potential effects of global warming on O3 and other photochemical oxidants, has renewed
23 interest in the effects of air pollution on both managed and unmanaged terrestrial ecosystems
24 (Adams et al., 1998). In addition, concern is growing for regarding the effects of air pollutants
25 on natural ecosystems and on the services they provide (Daily, 1997). Unfortunately, this
26 interest has not yet translated into additional peer-reviewed publications addressing O3 or other
27 air pollutants effects on ecosystems.
28 This section of the current criteria document first discusses the availability of economic
29 information and its usefulness in forming environmental policy. Next, economic assessments of
30 air pollution effects and findings from the 1996 AQCD (U.S. Environmental Protection Agency,
31 1996) are discussed, followed by a synthesis of the limited literature available since the 1996
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1 AQCD with respect to O3 effects on agriculture, forestry, and ecosystems. Finally, limitations
2 and continuing uncertainties are reviewed. The most fundamental of these is the lack of
3 measurements of the economic effects of air pollution on natural ecosystems. Other issues
4 include the variability of performance in both managed and natural ecosystems under increased
5 climatic and air pollution variability as well as the challenges related to spatial and temporal
6 scales used in performing economic assessments. To date, this set of effects has been sparsely
7 addressed.
8
9 9.8.2 The Measurement of Economic Information
10 Economic science is an exercise in deductive logic in which testable hypotheses about the
11 behavior of economic agents (i.e., farmers, consumers, resource owners) and markets are
12 deduced from a body of theory. That body of theory is based on a series of premises proposed
13 by economists and philosophers dating back over two centuries to Adam Smith. These premises
14 gradually evolved into a theoretical foundation primarily dealing with microeconomics,
15 culminating in structural relationships that define the operation of markets. This foundation was
16 first laid out in a comprehensive and rigorous fashion by Alfred Marshall in 1920. Paul
17 Samuelson (1948) formalized these theoretical relationships, resulting in what is sometimes
18 referred to as modern, or neoclassical, economics.
19 The insights gained from the theoretical foundation of economics helped shape the nature
20 of applied economic, or policy-relevant research. An example of such applied research is when
21 economists seek to measure the economic consequences of air pollution on agriculture. Such an
22 application is described in Adams and Horst (2003), who provided a graphical representation of
23 the measurement of the effects of air pollution on the well-being of producers and consumers.
24 Economic theory is applied to real-world problems when the methods of economics and the need
25 for data from other disciplines come into play. When measuring the economic effects of an
26 environmental change, economists need an economic model or method that is theoretically
27 consistent, i.e, defensible, as well as data with which to estimate both economic and
28 environmental science relationships for use in the model. Among accepted economic assessment
29 methods, the actual choice is frequently determined by the nature of the problem to be addressed.
30 It should be noted that the choice of assessment method can affect the type of economic
31 information that is obtained. Even with a given assessment method, results are sensitive to
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1 specific data treatment or assumptions (Adams, 1999). For example, some methods only
2 measure effects on a particular group e.g., farmers. Other methods may measure effects across
3 several groups. Thus, one should not expect the magnitudes of damage or benefits to be
4 identical across economic assessments. One should, however, expect that the direction of the
5 effects will be similar.
6 Once it is established that an assessment meets basic economic criteria, e.g., including
7 human behavioral responses, the selection of the specific economic assessment method is often a
8 relatively minor issue in terms of estimating benefits of air pollution control (or disadvantages of
9 increases in air pollution). Although results differ across approaches, the differences are largely
10 attributable to specific features of the assessment (e.g., whether the natural science data include a
11 particular effect or relationship, whether effects on consumers are measured, and so forth). The
12 nature and quality of the air quality forecasts used in the assessments can greatly influence the
13 sensitivity of the assessments (Adams and Crocker, 1989). This is particularly noticeable when
14 dealing with forecasts of seasonal air pollution changes (Adams et al., 1988). From the
15 standpoint of providing policy guidance, the differences in economic estimates attributable to the
16 assessment methods are often swamped by uncertainty in the natural and physical science
17 forecasts. This has also been noted in recent economic assessments of climate change (Adams
18 et al., 1998). In many settings, the quality of economic assessments of air pollution is likely to
19 be improved more by refining the physical and natural science data used in the assessments than
20 by intensive efforts to fine-tune the assessment techniques (Adams, 1999).
21
22 9.8.3 Understanding of Air Pollutants Effects on the Economic Valuation of
23 Agriculture and Other Vegetation in the 1996 Criteria Document
24 Evidence from the plant science literature cited in the 1996 O3 AQCD (U.S. Environmental
25 Protection Agency, 1996) is unambiguous with respect to the adverse effects of tropospheric O3
26 on some types of vegetation. For example, the 1996 AQCD noted that findings from the EPA
27 multiyear NCLAN program in the 1980s provided rigorous corroboration of at least two decades
28 of previous research and a century of anecdotal observations that showed that O3 at ambient
29 levels caused physical damage to plants in general and to important crop species in particular.
30 Specifically, NCLAN established that ambient O3 levels resulted in statistically significant
31 reductions in yields for some crop species (Heagle et al., 1988). The 1996 AQCD also assessed
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1 the results of studies regarding O3 effects on crops, forests, and natural vegetation in more detail.
2 More recent reviews, such as the comprehensive survey of the economic literature on
3 agricultural effects by Spash (1997) corroborate the synthesis of results reported in the 1996
4 AQCD.
5 The number and quality of assessments of the economic consequences of O3 exposures on
6 vegetation reported in the 1996 AQCD are primarily a function of the state of evidence obtained
7 from scientific studies in each vegetation category. For example, the plant science evidence
8 reviewed in the 1996 AQCD concerning effects of O3 exposures on agricultural crops was
9 reported to be more valid than for individual tree species or plant communities (ecosystems).
10 As a result, most economic assessments discussed in the 1996 AQCD focused on the data
11 obtained from studies of agricultural crops. The economic literature dealing with O3 effects on
12 forest productivity in the 1996 AQCD is sparse. The few economic assessments of tree or forest
13 effects reported in the 1996 AQCD were confined to evaluations of assumed or hypothetical
14 changes in output, such as board feet of lumber (e.g., Haynes and Adams [1992]). As noted in
15 the 1996 AQCD, O3 effects on ecosystems and their services had not been measured in any
16 systematic fashion and no peer-reviewed economic assessments were yet reported.
17 This section first briefly reviews economic assessments drawn from the review in the 1996
18 criteria document. This review is the benchmark against which recent articles are then discussed
19 in the subsequent section. As was the case in 1996, the discussion of economic valuation of
20 ecosystem effects is generally limited to conceptual and methodological issues, given the
21 continued lack of empirical analyses in this category.
22
23 9.8.3.1 Agriculture
24 In view of the importance of U.S. agriculture for both domestic and world consumption of
25 food and fiber, reductions in U.S. crop yields could adversely affect human welfare. The
26 plausibility of this premise has resulted in numerous attempts to assess, in monetary terms, the
27 losses from ambient O3 exposures or the benefits of O3 control, to agriculture. Twenty-three
28 assessments of the economic effects of O3 exposures on agriculture were reviewed in the 1996
29 AQCD, highlighting key issues in the validity of these assessments (U.S. Environmental
30 Protection Agency, 1996). First, the evidence should reflect how crop yields will respond under
31 actual field conditions to O3 exposures. Second, the air quality data used to frame current or
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1 hypothetical effects of O3 on crops should represent actual exposures sustained by crops at
2 individual sites or production areas. Finally, the assessment methodology into which such data
3 are entered should (1) capture the economic behavior of producers and consumers as they
4 adjust to changes in crop yields and prices that may accompany changes in O3 air quality;
5 (2) accurately reflect institutional considerations, such as regulatory programs and income
6 support policies (e.g., provisions of federal Farm Bill legislation), that may result in market
7 distortions; and (3) use measures of well-being that are consistent with economic principles.
8 Assessments of O3 damage to agricultural crops reported in the 1996 AQCD employed
9 procedures for calculating economic losses that met the conditions described above. More
10 specifically, the assessments provided 23 quantitative estimates of the economic consequences
11 of exposures to O3 and other air pollutants to agriculture that reflect producer-consumer
12 decision-making processes, associated market adjustments, and some measure of distributional
13 consequences between affected parties. Many of the economic assessments reviewed in
14 previous O3 documents also focused on O3 effects in specific regions, primarily California and
15 the Corn Belt (e.g., Garcia et al., 1986). This regional emphasis in the earlier literature may be
16 attributed to the relative abundance of data on crop response and air quality for selected U.S.
17 regions, as well as the importance of some agricultural regions (such as California) for the U.S.
18 agricultural economy.
19 Two U.S. national studies described in previous O3 criteria documents that are worthy of
20 additional comment are Kopp et al. (1985) and Adams et al. (1986). These were judged to be
21 "adequate" in terms of the three critical areas of data inputs in the 1996 AQCD. Together, it was
22 argued, they provide a reasonably comprehensive estimate of the economic consequences of
23 changes in ambient air O3 levels on agriculture. Because of their central role in the 1996 criteria
24 document, these two studies are reviewed briefly below.
25 The Kopp et al. (1985) and Adams et al. (1986) studies indicated that ambient levels of O3
26 were imposing substantial economic costs of ~$3.4 billion (in 2000 U.S. dollars) on agriculture.
27 Both were judged to suffer from several sources of uncertainty, but the document concluded that
28 these possible improvements in future assessments were not likely to greatly alter the range of
29 agricultural benefit estimates arising from O3 reductions for several reasons. First, the studies
30 covered about 75 to 80 % of U.S. agricultural crops (by value). For inclusion of the other 20%
31 to significantly change the estimates would require that their sensitivities to O3 be much greater
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1 than for the crops that have been included to date. Second, model-sensitivity analyses reported
2 in past studies indicate that changes in plant exposure response relationships must be substantial
3 to translate into major changes in economic benefits estimates. For example, it was assumed
4 unlikely that use of different exposure measures, or inclusion of interaction effects, would
5 greatly alter the magnitude of the economic estimates. Third, it was believed that countervailing
6 effects would mitigate against large swings in the estimates, e.g., longer O3-exposure periods
7 may predict greater yield losses, but O3-water stress interactions tend to reduce loss the
8 yield estimates.
9 Other national assessments reported in the 1996 AQCD provided general corroboration of
10 the results of Kopp et al. (1985) and Adams et al. (1986). An evaluation of these studies in
11 terms of the adequacy of information from plant exposure studies, and aerometric and economic
12 data was presented in the 1996 AQCD, along with estimates of benefits or damages associated
13 with changes in O3. Most of the studies built on either Kopp et al. (1985) or Adams et al. (1986).
14 A relevant question was whether subsequent studies provided any "surprises" in terms of
15 magnitude of economic effects.
16 Common themes or findings from these and earlier O3 and other air pollution studies were
17 summarized in two synthesis papers, Adams and Crocker (1989) and Segerson (1991). The
18 major conclusion is that the agricultural effects of tropospheric O3 at ambient levels impose
19 economic costs to society or, conversely, that reductions in ambient O3 should result in societal
20 benefits.
21 Several studies contained in the 1996 AQCD still are of interest. For example, one finding
22 pertains to the relationship between federal farm programs and air pollution regulations
23 (McGartland, 1987). In each case, the inclusion of farm programs in the economic models
24 resulted in modest reductions in the economic benefits of O3 control due to increased farm
25 program costs. As Segerson (1987) noted, however, it is not clear that these increased costs
26 should be charged against the potential benefits of an O3-regulatory standard, but rather,
27 considered as an additional cost associated with the inefficiencies of the farm program. It should
28 also be noted that the nature of federal farm programs was changed dramatically by Congress in
29 1996 in an attempt to reduce the federal government's role in agriculture. Although more recent
30 federal legislation, such as the 2002 Farm Bill (U.S. Congress, 2002), appears to be restoring the
31 federal government's role in the farming sector, this issue currently is not as important as
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1 suggested by earlier studies, due to the declining reliance on deficiency payments to farmers,
2 which tend to distort resource allocation.
3 Another national study (Adams et al., 1988) analyzed economic benefits under a regulatory
4 alternative involving a seasonal (crop growing season) O3-exposure index measured as a 12-h
5 mean, instead of hourly levels or percent changes from ambient as reported in earlier studies.
6 Specifically, a seasonal average of 50 ppb O3 (measured as a 12-h seasonal mean) with a
7 95% compliance level, is reported in Adams et al. (1988). The result (a $2.9 billion benefit in
8 2000 dollars) is similar to the assumed 25% reduction across all regions reported by Adams et al.
9 (1986). At least one study also combined environmental stressors (e.g., O3, UV-B radiation) in
10 performing economic assessments. Adams and Rowe (1990), using the same model as Adams
11 et al. (1986), reported that a 15% depletion of stratospheric O3 (resulting in a 13% increase in
12 tropospheric O3) would cause an economic loss of ~$1.4 billion in 2000 dollars attributed to the
13 tropospheric O3 increase. Reducing VOCs/NOx motor vehicle emissions by 10% would result in
14 a benefit of ~$0.3 billion, while a complete elimination of motor vehicle emissions would yield a
15 benefit of ~$3.4 billion (1990 dollars). The range of these numbers is consistent with values
16 reported in Adams et al. (1986), Kopp et al. (1985), and other national-level analyses, i.e.,
17 estimates of from $1.0 to 2.0 billion for reductions in ambient O3 of 25 to 40%.
18
19 9.8.3.2 Forests (Tree Species) and Natural Ecosystems
20 The long-term nature of air pollution effects on perennial species such as trees creates
21 challenges to plant scientists in attempts to sort out the effects of specific individual stressors
22 such as O3 from among the many other potential causal factors (Skelly, 1988). It also creates
23 problems in terms of measuring the impacts on direct economic value, such as reductions in
24 board-feet of lumber produced per unit of time.
25 Most of the literature in the 1996 AQCD dealing with forest species reported the effects of
26 O3 exposures in terms of foliar injury (Davis and Skelly, 1992; Freer-Smith and Taylor, 1992;
27 Simini et al., 1992; Taylor, Jr. and Hanson, 1992). This emphasis on foliar effects in the forest
28 effects literature (rather than marketable yield) is similar to the state of science for agricultural
29 crops prior to 1980. More recent studies address the effects of air pollutants on forest tree
30 species diversity (Bringmark and Bringmark, 1995; Vacek et al., 1999; Weiner et al., 1997).
31 However, such information is of limited use in economic assessments. The exception is in
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1 measuring the economic value of aesthetic changes in a forest stock, where changes in species
2 composition may affect recreational values (Crocker, 1985).
3 The data concerning changes in marketed output such as board-feet of lumber or changes
4 in growth rates in managed forests, the affects on the growth of almond, peach, apricot, pear and
5 plum trees in orchards cited in the 1996 document (U.S. Environmental Protection Agency,
6 1996) have not been quantified. In addition, the economic impact of reductions in growth of
7 seedlings of evergreen trees, e.g., slash pine, presented in the same document have not been
8 valued. The few studies which attempted to measure economic losses arising from exposures to
9 O3 or other pollutants circumvented the lack of plant science data by assuming (often arbitrary)
10 reductions in forest species growth or harvest rates (Adams, 1986; Callaway et al., 1986;
11 Crocker and Forster, 1986; Haynes and Adams, 1992). Although the economic estimates
12 reported in the 1996 AQCD are comparable to those reported for agriculture (e.g., $2.6 billion
13 for eastern Canada forests, $2.9 billion for eastern U.S. forests in 2000 dollars), the lack of yield
14 and/or growth effects data makes these studies only suggestive at best, of the economic
15 consequences of forest effects directly attributable to O3 exposures questionable. Recent
16 developments in forestry economic modeling capabilities, in support of climate change research,
17 have enhanced the ability to measure the effects of environmental stressors on this sector
18 (Adams et al.,1996; McCarl et al., 1998). However, these models need data on changes in either
19 timber production or growth rates, both of which are lacking for forest species under alternative
20 O3 levels.
21
22 9.8.4 Studies Since 1996 of Ozone Exposure Effects on the Economic Value of
23 Agriculture, Forests, and Ecosystems
24 Of the few current (post-1996) economic studies addressing agricultural effects, none offer
25 new insights of value in determining the economic cost of O3 exposures. These post-1996
26 studies used variants of the economic methods from earlier assessments and measure yield
27 changes from response functions arising from the NCLAN or similar data. For example, Kim
28 et al. (1998) used a mathematical programming model of the San Joaquin Valley agricultural
29 sector in California, combined with crop yield response functions, to assess the economic effects
30 of O3 on California crops. Their results showed net benefits from reductions in ambient O3
31 levels, a finding consistent with all previous economic assessments. In another study,
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1 Westenbarger and Frisvold (1994) measured agricultural exposures to O3 (and acid precipitation)
2 in the United States. Though not an economic analyses of the costs of ambient exposures, they
3 identified areas of the United States of greatest potential economic damage based on the
4 interface between regional pollution levels and the value of crop production in each region.
5 A study by Murphy et al. (1999) of the economic effects of tropospheric O3 on U.S.
6 agriculture is of note here, because it confirms the general magnitude of economic effects
7 reported by the two key studies performed a decade earlier (Kopp et al., 1985; Adams, 1986).
8 Specifically, Murphy et al. (1999) evaluated benefits to eight major crops associated with several
9 scenarios concerning the reduction or elimination of O3 precursor emissions from motor vehicles
10 in the United States. Their analysis reported a $2.8 to 5.8 billion (1990 dollars) benefit from
11 complete elimination of O3 exposures from all sources, i.e., ambient O3 reduced to a background
12 level assumed to be 0.025 to 0.027 ppm. While the analytical framework is similar to Adams
13 et al. (1986) in the use of NCLAN-based yield response functions and a mathematical
14 programming-based economic optimization model, the study is novel in its focus on the role of
15 motor vehicle emissions of VOCs/NOx in total anthropogenic O3 levels. The study is also
16 notable in its careful attention to federal farm program effects, particularly the deficiency
17 payment component.
18 In addition to these studies in peer-reviewed journals, a number of site-specific effects
19 studies have been performed, primarily by consulting companies for state public utility
20 commissions. Although perhaps of use to public utility commissioners concerned with effects
21 from single power plants or other localized sources, these regional studies generally contribute
22 little to the assessment of air pollution effects at the national level. Also, such reports are not
23 peer reviewed. Therefore, they are not discussed here.
24 There have been a number of recent studies of air pollutant effects on tree species in the
25 literature. Some have reported changes in total biomass and focused on European species
26 (Kurczynska et al., 1997). Other studies have assessed changes in composition of forest species
27 (biodiversity) or forest health due to exposure to air pollutants (Bringmark and Bringmark, 1995;
28 McLaughlin and Percy, 1999; Vacek et al., 1999). As noted previously, changes in forest
29 biomass and composition are more difficult to value than marketable products. However,
30 measures of forest composition or health have implications for an area of increasing policy
31 concern, that being the effect of air pollutants and other environmental stressors on unmanaged
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1 (natural) ecosystems and the services they provide (Goulder and Kennedy, 1997; Pimentel et al.,
2 1997). Considerable discussion has occurred among ecologists and economists as to the
3 appropriate means for valuing these services (Anderson, 1990; Carpenter and Dixon, 1985;
4 Common and Perrings, 1992). A number of conceptual articles have been published on this
5 issue in both economic and ecological journals (Bergstrom, 1990; Castle, 1993; Pearce, 1993;
6 Suter, II, 1990).
7 A continuing empirical challenge concerns the lack of information on how changes in
8 biodiversity affect ecosystem performance resulting from O3 stresses, and the problem of
9 establishing economic values for such changes (Cairnes and Lackey, 1992; Norton, 1987; Pimm,
10 1997; Polasky, 2001; Randall, 1991). As noted in the 1996 AQCD, and more recently by Daily
11 (1999, 2000) and Polasky (2001), there continues to be a lack of empirical studies that actually
12 assess the economic value of changes in biodiversity or in service flows due to any
13 environmental stressors. While some studies report monetary estimates, the estimates are
14 generally for expository purposes and those would not be as defensible as the agricultural studies
15 described earlier. For example, Costanza et al. (1997) assigned a value to the world's
16 ecosystems, but the procedures used render this an exploratory study at best. As assessed by
17 Polasky (2001), "In general, the field of valuation of ecosystem services is in its infancy."
18 He attributed the lack of empirical studies due to both a lack of the understanding of ecology of
19 ecosystem services and to the absence of reliable methods to estimate the value of these services.
20 In summary, the studies of crop and forest responses in the economic literature indicate
21 that O3 reduces crop yields and imposes economic costs. The economic literature also indicates
22 that O3 adversely influences the physiological performance of tree species and demonstrates, as
23 expected, that changes in growth have economic consequences. However, the economic data
24 and literature available on ecosystem effects of O3 exposures are not sufficient to determine the
25 economic costs.
26
27 9.8.5 Limitations of Scientific Studies and Economic Information
28 The 1996 O3 AQCD discussed the need for additional research on both ecological
29 functions and economic methodology in order to better understand the economic implications of
30 air pollutants on ecosystem services. As noted by Daily (1999, 2000), Polasky (2001), and
31 others, this research agenda continues to be important. Despite the large number of discussion
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1 and survey articles published since 1996, there still are not sufficient data by which to estimate
2 confidently the magnitude of economic effects of O3 to forests and natural ecosystems. Nor is it
3 apparent that ongoing research is adequate to answer this question in the next criteria document
4 cycle. Specifically, there does not appear to be any comprehensive, ecological studies underway
5 that attempt to measure changes in ecosystem outputs under alternative O3 or other air pollutant
6 levels. Thus, in the near term, ecosystem services can only be discussed in qualitative terms.
7 However plausible the likelihood of economic damages to ecosystems, the available scientific
8 and economic information does not provide specific guidance on the magnitudes of these effects.
9 Beside the need to improve our understanding of the effects of O3 exposures on natural
10 ecosystems, a number of areas of research could help assess the full economic consequences of
11 such pollutants on managed ecosystems. The first of these is the relationship between O3
12 exposure levels and the variability of crop yields or changes in forest biodiversity. Most
13 assessments are based on the average or expected yield response of a crop to air pollution
14 exposure. However, the variability in yields (the spread or dispersion around the mean) appears
15 to be affected by the nature of plant exposure to pollutants (Hogsett et al., 1985; Musselman
16 et al., 1983). Plants exposed to the same mean dose but with different second moments of the
17 distribution of exposure may have different mean yields (Altshuller and Lefohn, 1996; Lefohn
18 and Benedict, 1982). In addition, the variability of the yield of the plant may also be increased
19 by greater variability in exposures (e.g., a higher frequency of extreme events). The economic
20 significance of higher yield varieties is such that variability may impose additional economic
21 costs, because most farmers have been reported to be averse to risk and prefer less variability for
22 a given yield (or profit). To date, no economic assessments of O3 damages to agriculture or
23 vegetation in general include risk-averse behavior (studies cited here assumed risk neutrality).
24 To assess the economic consequences of a relationship between farmers' risk preferences and
25 O3-induced changes in yield variability will require more information on the potential effects of
26 changes in O3 on crop yield variability. While no economic studies were found on the effects of
27 O3-induced changes in yield variability, a number of recent studies of climate change effects on
28 crop yields have indicated increased economic costs in the presence of increased climatic
29 variability (e.g., Mearns et al., 1997; Dixon and Segerson, 1999). Analogous economic costs
30 would be expected for changes in air pollution distributions; and these effects need to be
31 examined and quantified.
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1 Another research area concerns the need improvement in our understanding of temporal
2 (dynamic) and spatial characteristics of O3 exposures and their implications for crop yields,
3 production and producer profits. Most economic studies are static, in that they compare two
4 states of the world (e.g., economic activity at one O3 level versus at an alternative level). In
5 addition, most national- level studies (the type needed to evaluate SNAAQS) display coarse
6 regional-level resolution in terms of crop response, O3 exposure, and economic behavior. The
7 responses of producers to changes in yields due to changes in O3 levels are generally assumed to
8 be similar over geographical areas up to several states in size. However, the changes between air
9 quality scenarios are more likely to be characterized by transient changes in exposure levels,
10 which means the producer responses are also likely to be gradual, rather than abrupt. Similarly,
11 the lack of finer scale (regional-level) data and modeling capabilities suggests that important
12 micro-level physical and economic effects are ignored. To what extent these abstractions
13 influence net economic effects is an empirical question. Research on these types of abstractions
14 and assumptions within other economic settings, such as climatic change, have shown that they
15 have implications for economic measurements (Adams, 2002).
16 Another issue is the natural or background level of O3 (or other pollutants of interest)
17 assumed in economic studies. While many economic studies focus on changes in pollution
18 levels from current conditions, some studies have measured the economic damages between an
19 assumed, or pristine, level and current levels in agricultural areas. Such an analysis, it is
20 reasoned, will provide a measure of the net damages due to anthropogenic sources. The
21 challenge here is to have a correct measure of the background level of the pollutant. Recent
22 research by Lefohn et al. (2001) has suggested that background levels may be considerably
23 higher than assumed in some of the previous economic assessments reported in the 1996 AQCD
24 (25 to 30 ppb in most studies). For example, Lefohn et al. (2001) detected hourly readings of
25 from 40 to 80 ppb during winter and spring months in remote areas of the United States.
26 If background O3 levels are in the range, then the economic damages estimated in studies with
27 lower background levels will be overstated. The issue of the background O3 level is important to
28 all assessments of vegetative damages due to O3.
29 In terms of expanding economic methods for future assessment, analysts should consider
30 using more "reduced form" estimation methods, particularly in situations where the availability
31 of dose-response functions is limited. This estimation approach is exemplified by Garcia et al.
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1 (1986). Specifically, their econometric study of the impact of O3 on producer profits used such a
2 reduced-form approach. In this approach, farmer actions are modeled as a function of ambient
3 O3 levels (along with other explanatory variables) without the direct use of dose response
4 functions. The advantage of this procedure is that one source of modeling uncertainty, the need
5 for dose-response functions, including time-consuming crop experiments to generate data, is
6 reduced. Actual responses of farmers' profits across air pollution gradients of ambient pollution
7 are observed instead of hypothesized. Although this procedure has not been widely used in air
8 pollution economic assessments, it has been used in a number of relatively recent climate change
9 studies (e.g., Mendelsohn et al. [1994]). The reduced-form method suffers from the fact that if
10 proposed O3 levels are lower than those observed in the estimation sample, then the prediction
11 accuracy of the method deteriorates. Also, some dose-response information is needed, if only to
12 establish the plausibility of the economic estimates.
13 Another area that may improve economic assessments is incorporating consideration of
14 livestock issues. To date, most agricultural economic assessments of O3 impacts have ignored
15 the livestock sector. Presumably this is because ambient O3 levels do not noticeably affect meat
16 yields. However, if feed prices and pasture conditions are affected by ambient O3 levels, then a
17 more accurate estimate of economic impacts would be forthcoming by including this link to
18 livestock in the assessment. These types of feed production and feed price effects are included
19 in the mathematical programming model used in Adams et al. (1986), but not in most other O3
20 effects assessments. The significance of livestock-feed linkages are demonstrated in a recent
21 study in the Netherlands by Kuik et al (2000). Using a mathematical programming model
22 similar to that in Adams et al. (1986), Kuik et al. (2000) found that livestock effects were
23 prominent, due mainly to improved pasture yields under reduced ambient O3 levels.
24
25 9.8.6 Conclusions
26 Substantial progress has been made over the past two decades in our understanding of the
27 effects of O3 and other oxidants on vegetation, particularly for agriculturally important plant
28 species. The physical and economic effects on agriculture are well documented and provide
29 useful information for the setting of SNAAQS. Effects on forests and natural ecosystems remain
30 problematic, due to limitations in natural science data and economic methods. The problem is
31 most acute for valuing natural ecosystem service flows.
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1 The current limitations surrounding forests and natural ecosystems present a rich research
2 agenda. However, not all research needs are likely to lead to better policies. Thus, areas of
3 greatest potential value in terms of regional policymaking need to be prioritized. Such priority
4 setting can be assisted by sensitivity analyses with existing economic models. By measuring the
5 changes in economic effects arising from changes in key parameters, research data gaps most
6 likely to affect economic values can be identified.
7
8
9 9.9 SUMMARY AND CONCLUSIONS FOR VEGETATION AND
10 ECOSYSTEM EFFECTS
11 9.9.1 Introduction
12 A significant number of ozone-related studies were published between 1996 and 2004, and
13 they are reviewed in this document in the context of earlier publications reviewed in the previous
14 criteria documents (U.S. Environmental Protection Agency, 1878; 1986; 1992; 1996).
15 In general, there has been a shift away from chamber studies in favor of more field-based
16 approaches although chamber exposures still dominate the effects literature. Field-based
17 approaches include increased survey's of visible injury as well as free-air exposure systems that
18 eliminate some of the problems associated with chamber effects. Increased emphasis has also
19 been placed on quantifying aspects of ozone uptake to better link ambient exposure monitoring
20 with plant/tree response. Much of the progress in quantifying uptake has occurred in Europe
21 with application of models simulating uptake. Evaluation of this new information has added to
22 our knowledge, but has not fundamentally altered the conclusions of previous Ozone Criteria
23 Document (U.S. Environmental Protection Agency, 1996).
24 It is known that ozone is phytotoxic, and that toxicity occurs only if ozone or its reaction
25 products reach the target tissues in the plant cell. Recent studies have provided an increased
26 understanding of how ozone interacts with the plant at the cellular level. Progress has been
27 made on understanding the initial stages of ozone toxicity, which resemble a "wound" response
28 in plants. The alteration of normal metabolism due to a "wound" starts in the cytoplasm, and
29 'spreads' to other cellular organelles. One of the secondary reactions is linked to an activation of
30 a senescence response that results in loss of the critical enzymatic components of photosynthesis.
31 The loss of photosynthetic capacity is in turn linked to the lowered productivity of plants and
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1 problems with efficient translocation of carbon. Although photosynthesis and translocation still
2 occurs at a reasonable rate, the loss of productivity, while better understood than in 1996, is still
3 not clearly explained. Dramatic strides in understanding the genetic make-up of plants, gene
4 control, and signal transduction/control over the last few years will accelerate in the future. This
5 understanding will translate into better models, more detailed schemes of how O3 alters much of
6 the basic metabolism of plants, and how to construct an index that better quantifies exposure and
7 effect. However, the translation of those mechanisms into how O3 is involved with the altered
8 cell metabolism and subsequent reductions in whole plant productivity and other physiological
9 facts has not yet been fully solved.
10 A number of biotic and abiotic factors are known, or suspected, to alter plant response to
11 ozone. New information published since 1996 does not significantly change the conclusions
12 arrived at in the previous criteria document regarding these multiple interactions. A major
13 concern from the last review was the potential, but unquantified, interaction of ozone and
14 diseases and insects. The nature of these interactions with O3 are probably dictated by the
15 unique, highly specific biochemical relationships between pathogen and host plant. At this time,
16 therefore, although some diseases may become more widespread or severe as a result of
17 exposure to O3, it is still not possible to predict which diseases are likely to present the greatest
18 risks to crops and forests.
19 Considerable emphasis during the last decade has been placed on research into O3
20 interactions with the components of global climate change: increased atmospheric CO2,
21 increased mean global temperatures, and increased surface level UV-B radiation. To date, the
22 limited experimental evidence and that obtained by computer simulation suggest that even
23 though an enriched CO2 atmosphere (-600 ppm) would more than offset the impact of O3 on
24 responses as varied as wheat yield or young Ponderosa growth, the concurrent increase in
25 temperature would reduce but probably not eliminate the net gain. A similar decrease in the net
26 gain resulting from the complete reversal by CO2 of the inhibition of photosynthesis caused by
27 O3 has been reported for increased UV-B irradiation. Overall, results are preliminary, based
28 upon minimal data, and do not allow adequate prediction of ozone effects under future climate
29 scenarios. Temperature is unquestionably also an important variable affecting plant response to
30 O3 in the presence of the elevated CO2 levels contributing to global climate change. In contrast,
31 evidence continues to accumulate to indicate that exposure to O3 sensitizes plants to low
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1 temperature stress, by reducing belowground carbohydrate reserves, which may lead to
2 responses in perennial species ranging from rapid demise to impaired growth in subsequent
3 seasons.
4 Efforts increased in the area of relating ozone flux (e.g., uptake) to response. Flux is
5 affected by a host of biotic and abiotic factors, including boundary layer conditions, stomatal
6 conductance, moisture conditions, solar irradiance and others. All of these factors provide a
7 more accurate estimation of flux for predicting plant, crop, or forest response. While this area of
8 research is promising, currently there is insufficient data linking flux with response across a
9 range of species or in different ecosystems. In addition, a flux-based standard would require
10 selection of a particular species or species assemblage across a meteorologically similar area in
11 order to adequately model the range of factors known to influence uptake. At this time, there is
12 insufficient information available to employ this approach nationally, but with continued
13 research may eventually provide an alternative index to the current exposure index.
14
15 9.9.2 Methodology
16 The majority of ozone effects studies are fumigation studies conducted in controlled
17 chambers, as noted in the previous document. The previous document noted that open-top
18 chambers (OTC's) represent the best technology for determination of crop yield to ozone at the
19 present time. While OTC's are still the best method for conducting controlled exposures of
20 varying length and frequency for exposure-response relationships, several new approaches have
21 been applied to ozone research, most notably free-air exposure or FACE systems. FACE
22 systems eliminate many of the concerns raised about closed or open-top chamber experiments,
23 including small plot size, altered microclimate within OTC's, and the effect of charcoal filtering
24 on overall air quality within OTC's. Although FACE systems have increased our understanding
25 in some areas, in most cases results from FACE systems have confirmed what we already
26 understood or hypothesized about how plants and plant assemblages respond to ozone.
27 In particular, studies with aspen showed that ozone symptom expression was generally similar in
28 OTC's, FACE, and also sites along an ambient ozone gradient, supporting the previously
29 observed variation among aspen clones using OTC's (Isebrands et al., 2000; 2001; Karnosky
30 et al., 1999). In addition, root growth is often found to be the most sensitive biomass response to
31 ozone, as reported in the previous document. Caution is still needed, however when making
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1 quantitative extrapolations from OTC's, and free-air exposure systems will continue to provide
2 valuable information in support of scaling efforts.
3 The lack of rural monitors continues to be a major problem in the characterization of ozone
4 exposures in remote areas, and in linking effects with exposure in natural ecosystems. Since the
5 last document, the use of passive samplers have expanded monitoring efforts to remote areas that
6 were previously uncharacterized. This has greatly enhanced our ability to link ozone
7 symptomology with elevated exposure in remote areas. Passive samplers do not capture
8 temporal dynamics of exposure, however; therefore, passive samplers are not a substitute for
9 active monitors when attempting to link dynamics of exposure with plant response, or when
10 developing exposure/dose response relationships that are needed in the standard setting process.
11 Exclusion methods such as EDU are the least disruptive of ambient culture conditions in
12 the field, as noted in the previous document. However, the level of protection afforded by EDU
13 is site and species specific, and is subject to local meteorologic conditions. Because of the
14 variability observed in the level of protection provided, and the fact that mechanisms of
15 protection afforded by EDU and other exclusion methods are unknown, caution is needed in
16 applying this approach to the study of ozone effects in the field.
17
18 9.9.3 Mode-of-Action
19 There are several steps in the process of ozone uptake and toxicity that are better
20 understood than in 1996 and are important in developing more refined hypotheses on ozone
21 uptake and mode of action on plants, and developing a flux-based index for use in quantifying
22 response, and ultimately are relevant for developing an air quality standard. Theses are listed
23 below:
24 (1) The entrance of O3 into the leaf through the stomata remains the critical step in O3
sensitivity. Not only does O3 modify the opening of the stomata, usually closing it
partially, but O3 also appears to alter the response of stomata to other stressful
situations, including a lowering of water potential and ABA responses. The
concentration of O3 within the leaf is not the same as the external concentration due
to reactions within the leaf but it is not "zero" (Moldau and Bichele, 2002).
25 (2) The initial reactions of O3 within the leaf are still unclear but the involvement of
hydrogen peroxide is clearly indicated. The detection of possible products by EPR
has progressed but is still not at the point where any products can be identified.
Nonetheless, reaction of O3 or its product with ascorbate and possibly other
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antioxidants present in the apoplastic space of the mesophyll cells is clear and serves to
lower the amount of O3 or product available to alter the plasma membrane of the cells.
26 (3) The initial sites of membrane reactions seems to be at the transport properties and
possibly the external signal transducer molecules. The alteration and mechanism of
the alteration of the varied carriers of K+and Ca2+ is far from clear but it would seem
that one of the primary triggers of O3-induced cell responses is due to a change in
internal Ca2+ levels (Cazale et al., 1998).
27 (4) The primary set of metabolic reactions that O3 triggers is now clearly the "wounding"
responses like that generated by cutting of the leaf or by pathogen/insect attack. This
seems to be due to a rise in cytoplasmic Ca2+. Ethylene release and alteration of
peroxidases and PAL activities as well as activation of many "wound"-derived genes
seem to be linked and some of the primary reactions to ozone exposure.
28 (5) The alteration of normal metabolism due to a "wound" has affects outside of the
cytoplasm. What effects are due to the "spreading of the problem" to other cellular
organelles is less clear. One of the secondary reactions is linked to an activation of a
senescence response. Certainly the loss of Rubisco and its messenger RNA is linked to
an early senescence or a speeding up of normal development leading to senescence.
Certainly the loss of photosynthetic capacity is linked to the lowered productivity of
plants and problems with efficient translocation are indicated. Although photosynthesis
and translocation still occurs in a reasonable rate, the loss of productivity is not
clearly explained.
29 The new information available on the mode of action of ozone is in part a result of
30 improved molecular tools for following rapid changes that occur within the leaf. Clearly there
31 are many changes that occur within hours or possibly days of the exposure to O3. However,
32 there is another effect due to O3 which takes longer to occur and tends to be most obvious under
33 exposure to low O3 concentrations for long periods. These effects have been linked to
34 senescence or some physiological response very closely linked to senescence (i.e., translocation,
35 reabsorption, allocation of nutrients and carbon).
36 Langebartels et al. (1997) have discussed "memory" or "carry-over effects" within the
37 plant's to explain sensitivity to frost in the winter after exposure to ozone in the summer. Others
38 have argued that it is nutrient status of the tree during the over-wintering phase of its life and
39 chronic exposure to ambient O3 (less severe with fewer peaks of very high levels) which induces
40 (1) mineral nutrient deficiency; (2) alterations of normal metabolism, including translocation and
41 allocation of carbohydrates and probably nitrogen; (3) disturbance of normal transpiration and
42 diurnal cycling, leading to water stress (Schmieden and Wild, 1995). While general nutrient
43 concentrations within the foliage may not occur, it may be that localized deficiencies do. This is
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1 hard to observe or prove without a great deal of work on all portions of a tree and without a
2 general hypothesis of what is occurring.
3 It is important to note that the dramatic strides in understanding the genetic make-up of
4 plants, gene control, and signal transduction/control over the last few years will only accelerate
5 in the future. That understanding will translate into better models of the hypotheses listed and
6 more detailed schemes of how O3 alters much of the basic metabolism of plants. Thus overall,
7 understanding of how O3 interacts with the plant at a cellular level has been likewise
8 dramatically improved. However, the translation of those mechanisms into how O3 is involved
9 with the altered cell metabolism and subsequent reductions in whole plant productivity and other
10 physiological facts has not yet been fully solved. As the understanding of wounding responses
11 of plants and more genome details and varied plant mutants become available, the cellular and
12 physiological responses of plants to O3 exposures is slowly becoming clearer. However, more
13 studies need to be made on a larger variety of species before this information can be
14 incorporated into indices of response and a protective standard.
15
16 9.9.4 Modification of Growth Response
17 There have been only a few studies reported since the 1996 document and none that
18 significantly change the conclusions of the previous criteria document with regard to factors
19 known to alter plant response to ozone.
20 In the area of biotic interactions, new evidence with regard to insect pests and diseases has
21 done little to remove the uncertainties noted in the previous criteria document (see Docherty
22 et al., 1997; and Fluckiger et al., 2002 for recent reviews). Most of the large number of such
23 interactions that may affect crops, forest trees and other natural vegetation have yet to be studied.
24 The trend suggested previously that O3 increases the likelihood and success of insect attack has
25 received some support from recent studies, but only with respect to chewing insects. With the
26 economically important group of sucking insects such as the aphids, no clear trends have been
27 revealed by the latest studies. Hence, although it seems likely that some insect problems could
28 increase as a result of increased O3 levels, we are still far from being able to predict the nature of
29 any particular O3-plant-insect interaction, its likelihood or its severity.
30 The situation is a little clearer with respect to interactions involving facultative,
31 necrotrophic plant pathogens, with O3 generally leading to increased disease. With obligate,
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1 biotrophic fungal, bacterial and nematode diseases, however, there are twice as many reports
2 indicating O3-induced inhibitions than enhancements. The frequent reports that infection by
3 obligate biotrophs reduces the severity of O3-induced foliar injury does not result in true
4 "protection" since there are negative effects on the host plant of the disease/?er se. With
5 obligate biotrophs, the nature of any interaction with O3 is probably dictated by the unique,
6 highly specific biochemical relationships between pathogen and host plant. At this time,
7 therefore, although some diseases may become more widespread or severe as a result of
8 exposure to O3, it is still not possible to predict which diseases are likely to present the greatest
9 risks to crops and forests.
10 Recent studies have done little to improve our understanding of the nature of interactions
11 between O3 and root symbionts. Several studies have indicated that the functioning of tree root
12 symbioses with mycorrhizae may be adversely affected by O3, but there is also evidence that the
13 presence of mycorrhizae may overcome root diseases stimulated by O3. There is also evidence
14 that O3 may encourage the spread of mycorrhizae to the roots of uninfected trees.
15 The few recent studies of the impact of O3 on intra-specific plant competition confirmed
16 that grasses frequently show greater resilience than other types of plants. In grass-legume
17 pastures, the leguminous species suffer greater growth inhibition (Johnson et al., 1996a;
18 Nussbaum et al., 2000a). And the suppression of Ponderosa pine seedling growth by blue wild-
19 rye grass was markedly increased by O3 (Andersen et al., 2001). However, we are far from
20 being able to predict the outcome of the impact of O3 on specific competitive situations such as
21 success!onal plant communities, or crop-weed interactions.
22 Although some recent field studies have indicated that O3 impact significantly increases
23 with increased ambient temperature, other studies have revealed little effect of temperature.
24 Light, a component of the plant's physical environment, is an essential 'resource' whose energy
25 content drives photosynthesis and CO2 assimilation. It has been suggested that increased light
26 intensity may increase the sensitivity to O3 of light-tolerant species, but decrease that of shade-
27 tolerant species, but this appears to be an over-simplification with many exceptions.
28 Temperature affects the rates of all physiological processes based on enzyme-catalysis and
29 diffusion, and each process and overall growth (the integral of all processes) has a distinct
30 optimal temperature range. It is important to note that a plants response to changes in
31 temperature will depend on whether it is growing near its optimum temperature for growth or
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1 near its maximum temperature (Rowland-Bamford, 2000). But temperature is unquestionably an
2 important variable affecting plant response to O3 in the presence of the elevated CO2 levels
3 contributing to global climate change. In contrast, evidence continues to accumulate to indicate
4 that exposure to O3 sensitizes plants to low temperature stress, by reducing belowground
5 carbohydrate reserves, which may lead to responses in perennial species ranging from rapid
6 demise to impaired growth in subsequent seasons (i.e., 'carry-over effects').
7 Although the relative humidity of the ambient air has generally been found to increase the
8 adverse effects of O3 by increasing stomatal conductance and thereby increasing O3 flux, there is
9 abundant evidence that the ready availability of soil moisture results in greater sensitivity to O3.
10 The partial 'protection' against the adverse effects of O3 afforded by drought has been observed
11 in field experiments and modeled in computer simulations (Broadmeadow and Jackson, 2000).
12 There is also compelling evidence that O3 can predispose plants to drought stress (Maier-
13 Maercker, 1998). Hence the response will depend to some extent upon the sequence in which
14 the stresses occur, but, even though the nature of the responses is largely species-specific,
15 successful applications of model simulations have shown the way to larger scale predictions of
16 the consequences of O3 x drought interactions. However, it must be recognized that regardless
17 of the interaction, the net result on growth in the short-term is negative, although in the case of
18 tree species, other responses such as increased water use efficiency could be a benefit to survival
19 in the long term.
20 Somewhat analogously with temperature, it appears that any shift away from the nutritional
21 optimum may lead to greater ozone sensitivity, but the shift would have to be substantial before
22 a significant effect on response to O3 was observed. Mineral nutrients in the soil, other gaseous
23 air pollutants, and agricultural chemicals constitute chemical factors in the environment. The
24 evidence regarding interactions with specific nutrients is still contradictory: there is some
25 experimental evidence that low general fertility increases sensitivity to O3, while simulation
26 modeling of trees suggests that nutrient deficiency and O3 act less than additively, but there are
27 too many example of contrary trends to permit any sweeping conclusions.
28 Interactions of O3 with other air pollutants have received relatively little recent attention
29 since 1996 (see Barnes and Wellburn, 1998; and Fangmeier et al., 2002 for recent reviews).
30 The situation with SO2 remains inconsistent, but seems unlikely to pose any additional risk to
31 those related to the individual pollutants. With the NO and NO2 the situation is complicated by
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1 their nutritional value as a source of N. In leguminous species, it appears that NO2 may reduce
2 the impact of O3 on growth, with the reverse in other species, but the nature of the exposure
3 pattern, i.e., sequential or concurrent, also determines the outcome. Much more investigation is
4 needed before we will be able to predict the outcomes of different O3-NO-NO2 scenarios. The
5 latest research into O3 * acid rain interactions has confirmed that, at realistic acidities, significant
6 interactions are unlikely. A continuing lack of information precludes offering any
7 generalizations about interactive effects of O3 with NH3, HF, or heavy metals. More evidence
8 has been reported for protective effects against O3 afforded by the application of fungicides.
9 Considerable emphasis during the last decade has been placed on research into O3
10 interactions with the components of global climate change: increased atmospheric CO2,
11 increased mean global temperatures, and increased surface level UV-B radiation. However, it
12 must be noted that most of these studies have tended to regard increased CO2 levels and
13 increased mean temperatures as unrelated phenomena. Experiments into the effects of doubled
14 CO2 levels at today's mean ambient temperatures are of questionable value in trying to assess the
15 impact of climate change on responses to O3. To date, the limited experimental evidence and
16 that obtained by computer simulation suggest that even though an enriched CO2 atmosphere
17 (-600 ppm) would more than offset the impact of O3 on responses as varied as wheat yield or
18 young Ponderosa growth, the concurrent increase in temperature would reduce but probably not
19 eliminate the net gain. A similar decrease in the net gain resulting from the complete reversal by
20 CO2 of the inhibition of photosynthesis caused by O3 has been reported for increased UV-B
21 irradiation. However, these are preliminary results based upon minimal data.
22
23 9.9.5 Exposure Indices
24 The previous Criteria Document (U.S. Environmental Protection Agency, 1996a) noted
25 that the complexities associated with uptake and ozone interactions with external physical and
26 internal genetic factors that influence plant response makes the development of an 'ideal'
27 exposure index that characterizes plant exposure and response extremely difficult. As a result, a
28 biologically relevant surrogate for uptake was recommended that related ambient exposure to
29 measured growth/yield response. Despite additional research linking estimates of flux with plant
30 response since 1996, there is still insufficient information to identify a flux-based model that
31 incorporates sufficient complexity across space and time to be non-site or non-species specific.
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1 Therefore, based on the current state of knowledge, exposure indices that cumulate and
2 differentially weight the higher hourly average concentrations, but include the mid-level values,
3 still represents the best approach in the United States for relating vegetation effects to exposure.
4 The few studies published in this interim have substantiated earlier conclusions on the role
5 of exposure components, including concentration, duration and exposure patterns, in determining
6 the growth response of plants to ozone (Yun and Laurence, 1999; Oksanen and Holopainen,
7 2001). Recent studies using different exposure patterns have confirmed earlier studies on the
8 role of higher concentrations and exposure duration (Nussbaum et. al., 1995). A role for peak
9 concentrations is inferred based on improved air quality in regions like the San Bernardino
10 Mountains in southern California. The reductions of O3 in the San Bernardino area are
11 associated with reductions in the higher hourly average concentrations, but not in the midrange
12 hourly average concentrations, which are increasing (Lee et al., 2003; Lefohn and Shadwick,
13 2000). General forest improvement is reported following a decrease of O3 along a decreasing
14 gradient of exposure (Miller and Rechel, 1999; Arbaugh et al., 2003; Tingey et al., 2004). These
15 studies provide the basis for focusing on higher O3 concentrations, while including the lower
16 levels, when estimating the effects of emission reductions on vegetation.
17 New studies have demonstrated the potential disconnection of peak events and maximal
18 stomatal conductance at xeric to mesic sites in California (Panek et al., 2002; Panek, 2004;
19 Grulke et al., 2002). In addition, a few studies have indicated the uptake of ozone during
20 nighttime hours is greater than previously thought (Grulke et al., 2003; Massman, 2004), and a
21 review of the literature suggests a large number of species exhibit some degree of conductance at
22 night (Musselman and Minnick, 2000). These studies suggest a reconsideration of the need to
23 cumulate exposure 24 hrs per day and not just during daylight hours in exposure index. This
24 lack of coincidence in temporal patterns of conductance and peak ambient concentrations
25 introduces uncertainty in assessing ozone's impact. The use of an exposure index that does not
26 consider regionally unique climate and site factors which modify stomatal conductance may, as a
27 result, under or over estimate growth effects. The shortcomings of an ambient exposure-based
28 index is especially apparent when assessing ozone's potential impact across broad climatic
29 regions of the U.S. or Europe.
30 It is apparent that additional research is needed to develop indices which are more
31 physiologically- and meteorologically-connected to the actual dose of ozone the plant receives.
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1 The cumulative, concentration-weighted exposure indices are acknowledged surrogates for
2 effective dose that are conceptually simple and easy to measure. As discussed in the previous
3 Criteria Document, they do not fully characterize the potential for plant uptake and subsequent
4 response since they do not include the physical, biological, and meteorological processes
5 controlling the transfer of O3 from the atmosphere through the leaf and into the leaf interior.
6 The flux-based approach should provide an opportunity to improve upon the
7 concentration-based (i.e., exposure indices) approach. A great deal of progress has occurred in
8 development and testing of stomatal models that may be generally applicable across certain
9 vegetation types (e.g., Matyssek et al., 2004; Grunhage and Jager 2003; Danielsson et al., 2003;
10 Emberson et al., 2000; Pleijel et al., 2000). While a flux-based approach is preferred, a
11 cautionary argument was advanced in a few publications based on the non-linear relationship
12 between ozone uptake and foliar injury (growth not assessed). The concern is that not all O3
13 stomatal uptake results in a reduction in yield, which depends to some degree on the amount of
14 internal detoxification occurring with each particular species. Those species having high
15 amounts of detoxification potential may show little relationship between O3 stomatal uptake and
16 plant response (Musselman and Massman, 1999).
17 The European approach and acceptance of flux-based critical values is in part a recognition
18 of the problems associated with ambient exposure-based indices. Research continues across
19 Europe to develop the necessary experimental database and modeling tools that will be required
20 to provide the scientific basis for a critical level for ozone (Grunhage et al., 2004; Fuhrer et al.,
21 1997; Grunhage and Jager, 1994).
22 Given the current state of knowledge and the best available data, exposure indices that
23 cumulate and differentially weight the higher hourly average concentrations, but include the
24 mid-level values, continue to offer the most defensible approach for vegetation protection.
25 A large database exists that has been used for establishing exposure-response relationships, and
26 at this time, such a database does not exist for relating ozone flux to growth response.
27 It is anticipated that as the overlapping relationships of conductance, concentration, and
28 defense mechanisms are better defined, the flux-based indices may be able to predict vegetation
29 injury and/or damage with more accuracy than the exposure-response models. However, it is
30 unclear that such is the case at this time. The translation of these research and assessment tools
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1 to air quality standards has the additional need to be simple, understandable, and adaptive to a
2 manageable monitoring program.
3
4 9.9.6 Ozone Exposure-Plant Response Relationships
5 Data published since 1996 continue to support the conclusions of previous criteria
6 documents that there is strong evidence that ambient concentrations of O3 cause foliar injury and
7 growth and yield damage to numerous common and economically valuable plant and tree
8 species. For annual vegetation, the data summarized in Table 4-19 show a range of growth and
9 yield responses both within species and among species. Nearly all of these data were derived
10 from studies in OTCs, with only two studies using open-air systems in the UK (Ollerenshaw
11 et al., 1999; Ollerenshaw and Lyons, 1999). It continues to be difficult to compare studies that
12 report O3 exposure using different indices, such as AOT40, SUM06, or 7-h or 12-h mean values.
13 However, when such comparisons can be made, the results of recent research confirm earlier
14 results summarized in the previous criteria document (U.S. Environmental Protection Agency,
15 1996). A summary of earlier literature concluded that a 7-h 3-month mean of 49 ppb
16 corresponding to a SUM06 exposure of 26 ppm-h, would cause 10% loss in 50% of 49
17 experimental cases (Tingey et al., 1991). Recent data summarized in Table 4-19 support this
18 conclusion, and more generally indicate that ambient ozone exposures can reduce the growth and
19 yield of annual species. Some annual species such as soybean are more sensitive, and greater
20 losses may be expected (Table 4-19). Thus the recent scientific literature supports the
21 conclusions of the previous criteria document that ambient O3 concentrations are reducing the
22 yield of major crops in the United States.
23 Much research in Europe has used the AOT40 exposure statistic, and substantial effort has
24 gone into developing Level-1 critical levels for vegetation using this index. Based on regression
25 analysis of 15 OTC studies of spring wheat including one study from the U.S. and 14 from
26 locations ranging from southern Sweden to Switzerland, an AOT40 value of 5.7 ppm-h was
27 found to correspond to a 10% yield loss, and a value of 2.8 ppm-h corresponded to a 5% yield
28 loss (Fuhrer et al., 1997). Because a 4 to 5% decrease could be detected with a confidence level
29 of 99%, a critical level of an AOT40 value of 3 ppm-h was selected in 1996 (Karenlampi and
30 Skarby, 1996).
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1 In addition to reductions in crop yield, O3 may also reduce the quality or nutritive value of
2 annual species. Many studies have shown effects of O3 on various measures of plant organs that
3 affect quality, with most studies focusing on characteristics important for food or fodder. These
4 studies indicate that there may be economically important effects of ambient O3 on the quality of
5 crop and forage species. Previous criteria documents have concluded that visible symptoms on
6 marketable portions of crops and ornamental plants can occur with seasonal 7-h mean O3
7 exposures of 40 to 100 ppb (U.S. Environmental Protection Agency, 1978; 1986; 1996). The
8 recent scientific literature does not refute this conclusion.
9 The use of OTCs may reverse the usual vertical gradient in O3 that occurs within a few
10 meters above the ground surface (Section 4.2). This reversal suggests that OTC studies may
11 overestimate to some degree the effects of an O3 concentration measured several meters above
12 the ground. However such considerations do not invalidate the conclusion of the previous
13 criteria document that ambient O3 exposures are sufficient to reduce the yield of major crops in
14 the United States.
15 As for single-season agricultural crops, yields of multiple-year forage crops are reduced at
16 ozone exposures that occur over large areas of the U.S. This result is similar to that reported in
17 the previous criteria document (U.S. Environmental Protection Agency, 1996). When species
18 are grown in mixtures, O3 exposure can increase the growth of tolerant species while
19 exacerbating the growth decrease of O3-sensitive species. Because of this competitive
20 interaction, the total growth of the mixed-species community may not be affected by O3
21 exposure. However, in some cases mixtures of grasses and clover species have shown
22 significant decreases in total biomass growth in response to O3 exposure in studies in the U.S.
23 and in Sweden. In Europe, a provisional critical level for herbaceous perennials of an AOT40
24 value of 7 ppm-h over six months has been proposed to protect sensitive plant species from
25 adverse effects of O3.
26 For deciduous tree species, recent evidence from free air exposure systems and OTCs
27 supports results observed previously in OTC studies. For example, a series of studies was
28 undertaken using free air O3 enrichment in Rhinelander, WI (Isebrands et al., 2000; 2001).
29 These studies showed that O3 symptom expression was generally similar in OTCs, FACE, and
30 also sites along an ambient O3 gradient, supporting the previously observed variation among
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1 aspen clones using OTCs (Karnosky et al., 1999). As has been observed in previous criteria
2 documents, root growth often is found to be the most sensitive biomass response to O3.
3 Results since 1996 support the conclusion of the previous criteria document (U.S.
4 Environmental Protection Agency, 1996) that deciduous trees are generally less sensitive to
5 O3 than are most annual plants, with the exception of a few very sensitive genera such as
6 Populus and sensitive species such as black cherry. However, the data presented in Table 4-20
7 suggest that ambient exposures that occur in the U.S. can sometimes reduce the growth of
8 seedlings of deciduous species. Results from multiple-year studies sometimes show a pattern of
9 increased effects in subsequent years. In some cases however, growth decreases due to O3 may
10 become less significant or even disappear over time. While some mature trees show greater O3
11 sensitivity than do seedlings in physiological parameters such as net photosynthetic rate, these
12 effects may not translate into measurable reductions in biomass growth. However, because even
13 multiple-year experiments do not expose trees to O3 for more than a small fraction of their life
14 span, and because competition may in some cases exacerbate the effects of O3 on individual
15 species, determining effects on mature trees remains a significant challenge.
16 In Europe, a Level I critical level has been set for forest trees based on OTC studies of
17 European beech seedlings. A critical level was defined as an AOT40 value of 10 ppm-h for
18 daylight hours for a 6-month growing season (Karenlampi and Skarby, 1996). However, other
19 studies show that other species such as silver birch may be more sensitive to O3 than is beech
20 (Paakkonen et al., 1996).
21 For evergreen tree species as for other tree species, the O3 sensitivity of different genotypes
22 and different species varies widely. Based on studies with seedlings in OTCs, major species in
23 the US are generally less sensitive than are most deciduous trees, and slower growing species are
24 less sensitive than are faster growing species. There is evidence that interacting stresses such as
25 competition may increase the sensitivity of trees to O3. As for deciduous species, most
26 experiments with evergreen species have only covered a very small portion of the life span of a
27 tree, and have been conducted with seedlings, so estimating effects on mature trees is difficult.
28 For all types of perennial vegetation, cumulative effects over more than one growing
29 season may be important, and studies for only a single season may underestimate effects.
30 Mature trees may be more or less sensitive to O3 than are seedlings depending on the species,
31 but information on physiological traits can be used to predict such differences in some cases.
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1 In some cases, mature trees may be more sensitive to O3 than seedlings due to differences in gas
2 exchange rates, differences in growth rates, greater cumulative exposure, or the interaction of
3 ozone with other stresses.
4
5 9.9.7 Ecosystem Effects
6 There is evidence that tropospheric O3 is an important stressor of ecosystems, with
7 documented impacts on the biotic condition, ecological processes, and chemical/physical nature
8 of natural ecosystems. In turn, the effects of O3 on individual plants and processes are scaled up
9 through the ecosystem affecting processes such as energy and material flow, inter- and intra-
10 species competition, and NPP. Thus, effects on individual keystone species and their associated
11 microflora and fauna, which has been shown experimentally, may cascade through the
12 ecosystem to the landscape level, although it has not yet been demonstrated. By affecting water
13 balance, cold hardiness, tolerance to wind, and by predisposing plants to insect and disease pests,
14 O3 may even impact the occurrence and impact of natural disturbance (e.g., fire, erosion).
15 Despite the probable occurrence of such effects, there are essentially no instances where
16 ecosystem level, highly integrated studies have conclusively shown that ozone is indeed altering
17 ecosystem structure and/or function.
18 Systematic injury surveys demonstrate that foliar injury occurs on sensitive species in
19 many regions of the USA. However, the frequent lack of correspondence between foliar
20 symptoms and growth effects means that other methods must be used to estimate the regional
21 effects of O3 on tree growth rates. Investigations of the radial growth of mature trees in
22 combination with data from many controlled studies with seedlings and a few studies with
23 mature trees suggest that ambient O3 is reducing the growth of mature trees in some locations.
24 Studies using models based on tree physiology and forest stand dynamics suggest that modest
25 effects of O3 on growth may accumulate over time, and may interact with other stresses. For
26 mixed-species stands, such models predict that overall stand growth rate is generally not likely to
27 be affected. However, competitive interactions among species may change as a result of growth
28 reductions of sensitive species. These results suggest that O3 exposure over decades may be
29 altering the species composition of forests in some regions.
30 Despite increased understanding of possible ecosystem effects of ozone, the data base
31 demonstrating and quantifying the degree to which O3 is altering natural ecosystems is lacking.
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1 Much of the speculation of ozone impact on ecosystems must be inferred from a number of case
2 studies of forest plot field-based data reporting on a number of different species. One means to
3 discuss our current knowledge is by listing the areas in which there is lack of knowledge. These
4 include:
5 Ecosystem processes. Very little is known about the effects of O3 on water, carbon, and
6 nutrient cycling, particularly at the stand and community levels. Effects on belowground
7 ecosystem processes in response to O3, exposure alone and in combination with other stressors
8 are critical to projections at the watershed and landscape scales. Little is yet known about the
9 effects of O3 on structural or functional components of soil food webs or how these impacts
10 could affect plant species diversity (Andersen, 2003).
11 Biodiversity and genetic diversity. The study of genetic aspects of O3 impacts on natural
12 ecosystems has been largely based on correlations and it remains to be shown conclusively
13 whether O3 affects biodiversity or genetic diversity (Pitelka, 1988; Winner et al., 1991; Davison
14 & Barnes, 1998). Studies of competitive interactions under elevated O3 are needed (Laurence &
15 Andersen, 2003), and re-examination via new sampling of population studies to bring in a time
16 component to previous studies showing spatial variability in population responses to O3 are
17 needed. These studies could be strengthened by modern molecular methods to quantify impacts
18 on diversity.
19 Natural ecosystem interactions with the atmosphere. Little is known about feedbacks
20 between O3 and climate change on production of volatile organic compounds, which in turn,
21 could affect O3 production (Fuentes et al., 2001). At moderate to high O3 exposure sites,
22 aberrations in stomatal behavior could significantly affect individual tree water balance of
23 sensitive trees, and if the sensitive tree species is dominant, hydrologic balance at the watershed
24 and landscape level could be affected. This has not been addressed in any model because O3
25 exposure effects, if included at all in the modeling effort, have assumed a linear relationship
26 between assimilation and stomatal conductance. Interaction studies with other components of
27 global change (i.e., warming, increasing atmospheric CO2, N deposition, etc.) or with various
28 biotic stressors are needed to better predict complex interactions likely in the future (Laurence &
29 Andersen, 2003). Whether O3 will negate the positive effects of an elevated CO2 environment
30 on plant carbon and water balances is not yet known nor is it known if these effects will scale-up
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1 through the ecosystem. How O3 affects the progress of pest epidemics and insect outbreaks as
2 concentrations increase is unclear (Skarby et al., 1998).
3 Information concerning the impact of O3 on reproductive processes and reproductive
4 development under realistic field or forest conditions are needed as well as examination of
5 reproductive effects under interacting pollutants (Black et al., 2000).
6 Scaling. The vast maj ority of O3 studies of trees have been conducted with young,
7 immature trees and in trees that have not yet formed a closed canopy. Questions remain as to the
8 comparability of O3 effects on juvenile and mature trees and on trees grown in the open versus
9 those in a closed forest canopy in a competitive environment (Chappelka & Samuelson, 1998;
10 Kolb & Matyssek, 2001; Samuelson & Kelly, 2001). Scaling the effects of O3 across spatial
11 scales is also difficult. Scaling responses of single or a few plants to effects on communities and
12 ecosystems are complicated matters that will require a combination of manipulative experiments
13 with model ecosystems, community and ecosystem studies along natural O3 gradients, and
14 extensive modeling efforts to project landscape level, regional, national and international
15 impacts of O3. Linking these various studies via impacts on common research quantification
16 across various scales using measures of such factors as leaf area index or spectral reflective data,
17 which could eventually be remotely sensed (Kraft et al., 1996; Panek et al., 2003), would provide
18 powerful new tools for ecologists.
19 Identifying endpoints. In general, methodologies to determine the important values of
20 services and benefits derived from natural ecosystems are lacking. Identifying and quantifying
21 factors that could be used in comprehensive risk assessment for O3 effects on natural ecosystems
22 would increase societal awareness of the importance of protecting ecosystems (Heck et al.,
23 1998).
24
25 9.9.8 Economics
26 Substantial progress has been made over the past two decades in our understanding of the
27 effects of ozone and other oxidants on vegetation, particularly for agriculturally-important plant
28 species. The physical and economic effects on agriculture are well documented and provide
29 useful information for the consideration of establishing air quality standards for crops. Effects
30 on forests and natural ecosystems remain problematic, due to limitations in biological response
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1 data and economic methods. The problem is even more acute for valuing natural ecosystem
2 service flows.
3 The current limitations surrounding forests and natural ecosystems present a rich research
4 agenda. However, not all research needs are likely to lead to better policies. Thus, areas of
5 greatest potential value in terms of regional policy making need to be prioritized. Such
6 priority-setting can be assisted by sensitivity analyses with existing economic models; by
7 measuring the changes in economic effects arising from changes in key parameters, it is possible
8 to identify those research data gaps most likely to affect economic values.
9
10
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20
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i 10. TROPOSPHERIC OZONE EFFECTS ON UV-B FLUX
2 AND CLIMATE CHANGE PROCESSES
3
4
5 10.1 INTRODUCTION
6 In addition to exerting direct effects on human health and vegetation/ecosystems,
7 tropospheric ozone (O3) is involved in determining ground-level flux of solar ultraviolet (UV)
8 radiation and, also, in other processes that alter the earth's radiative balance and contribute to
9 resulting climate change. This chapter first discusses the role of tropospheric ozone in
10 determining surface-level UV flux and, then, secondly, discusses tropospheric O3 involvement in
11 global climate change.
12
13
14 10.2 THE ROLE OF TROPOSPHERIC OZONE IN DETERMINING
15 GROUND-LEVEL UV-B FLUX
16 Stratospheric O3 plays a crucial role in reducing the exposure of living organisms to solar
17 ultraviolet radiation. The stratosphere, the atmospheric layer immediately above the
18 troposphere, is roughly 50 km thick and possesses 90% or more of the atmosphere's O3.
19 Ozone depletion due to the release of long-lived anthropogenic chlorinated and fluorinated
20 hydrocarbons was discovered over the course of the 1970s and early 1980s and led to
21 establishment of an international treaty for the protection of stratospheric O3, the 1987 Montreal
22 Protocol on Substances that Deplete the Ozone Layer.
23 Active research has been underway into the effects of increased UV radiation on
24 ecosystems and on human health since the discovery of the seasonal polar "Ozone Hole" and the
25 declining stratospheric O3 concentrations in the midlatitudes. An outcome of this effort is a body
26 of literature that also describes the effects of tropospheric pollutants, PM, and O3, on ground-
27 level UV radiation flux. The Montreal Protocol requires routine review of the latest scientific
28 information available on the status of the O3 layer and of UV radiation levels at the earth's
29 surface. The World Meteorology Organization (WMO) and UN Environmental Program
30 (UNEP) are responsible for assessing the state of the science regarding the O3 layer and for
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1 reporting on trends in surface UV radiation levels. The latest WMO/UNEP assessment was
2 published in 2003 (WMO/UNEP, 2002).
3 This section describes the current level of scientific understanding of the factors
4 influencing the flux of UV radiation at the earth's surface, including geophysical factors,
5 tropospheric O3, PM, and cloud cover, with reference to the WMO/UNEP assessment and the
6 peer-reviewed literature. Factors influencing the degree of human exposure to UV-B and the
7 resulting health effects are also discussed.
8
9 10.2.1 Factors Governing Ultraviolet Radiation Flux at the Earth's Surface
10 10.2.1.1 UV Radiation: Wavelengths, Energies and Depth of Atmospheric Penetration
11 Designations for portions of the electromagnetic spectrum have evolved over time and are
12 usually associated with general function or effects caused by photons in a given wavelength
13 range. Gamma rays, at or below 0.1 nm in wavelength, are especially damaging high energy
14 photons emitted during radioactive decay and by stellar activity. Radiowaves, wavelengths
15 greater than 108 nm, are very low in energy and function as carriers for broadcast
16 communications. The energy possessed by any photon is inversely proportional to its
17 wavelength.
18 The wavelengths ranging between 50 and 400 nm in length are denoted "ultraviolet."
19 Solar radiation of wavelengths < 280 nm, including UV-C (200 to 280 nm), is almost entirely
20 blocked by Earth's upper atmosphere due to photoionization and photo-dissociation processes.
21 Figure 10-1 shows a comparison between solar flux above the atmosphere and the flux at the
22 earth's surface. The flux of solar radiation between 280 and 320 nm (UV-B) is absorbed or
23 scattered in part within the atmosphere, while radiation between 320 and 400 nm (UV-A) can be
24 scattered but are not absorbed by gases to any meaningful degree. UV-B and UV-A photons
25 contain the necessary levels of energy needed to break (photolyze) chemical bonds. Both UV-A
26 and UV-B are associated with human health- and ecosystem-damaging effects. Because UV-B
27 is more energetic, it is potentially capable of producing more biological damage than UV-A.
28 UV-A comprises 90% of the UV radiation that reaches the earth's surface, yet is 1000 times less
29 damaging to keratinocytes (skin cells), than UV-B (Hildesheim, 2004).
30
31
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0)
o
1.0000
I 0.1000
es
E
x 0.0100
0.0010
0.0001
Solar Flux
290
300 310 320
Wavelength (nm)
330
Figure 10-1. Extraterrestrial solar flux measured by the satellite UARS SOLSTICE
instrument (dotted line). The dashed line represents calculated atmospheric
transmittance and the solid line is the calculated absolute flux of UV
radiation for a solar zenith angle of SOdeg, total column O3 of 275 DU, and
a surface reflectivity of 8%. The fine structure on the surface flux trace
results from Fraunhofer lines (absorption specific wavelengths within the
solar atmosphere).
Source: Krotkovetal. (1998).
1 10.2.1.2 Temporal Variations in Solar Flux
2 The magnitude of the solar radiation flux entering the atmosphere is dependent upon long-
3 term solar activity, sunspot cycle (11 years), solar rotation (27 days) and the position of the earth
4 in its orbit around the sun. A variety of changes in solar irradiance can be found in the historical
5 data, beginning in 1700 and leading up to the present. Fligge and Solanki (2000) concluded that
6 solar irradiation changes on time scales of days to centuries can be attributed to variations in
7 solar magnetic features. Since the last Maunder minimum in 1700, solar irradiance has
8 increased very slightly, at approximately 3.0% for wavelengths < 300 nm and at -1.3% for the
9 UV-B and UV-A range. Including visible wavelengths, Fligge and Solanki (2000) estimated that
10 the overall increase in solar irradiance was -0.3%. Rozema et al. (2001) pointed out that any
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1 increase in wavelengths < 300 nm (UV-C) would initiate additional O3 formation in the
2 stratosphere, therefore increasing its UV radiation absorptive capacity.
3 Solar rotation and sunspot activity have the greatest effect on the radiation flux that
4 originates in the highest levels of the solar atmosphere. The amplitude of the associated cyclical
5 changes in solar shortwave radiation flux follows an inverse relationship between the photon's
6 wavelength and the solar altitude at which it was emitted. The maximum level of radiation
7 (solar-max) differs from the minimum (solar-min) by as much as 10% for wavelengths near
8 160 nm. This peak-to-trough difference declines to around 1% for 300 nm (UV-B range)
9 (Salby, 1996).
10 The combined effects of the earth's obliquity (the angle of the earth's axis of rotation with
11 respect to the plane of its orbit around the sun) and its precession (the rotation of the earth's axis
12 with respect to a perpendicular line through the plane of its solar orbit) yield variations of up to
13 30% in total summertime solar flux, depending on latitude (Hartmann, 1994).
14
15 Zenith Angle: Latitude, Season and Time of Day
16 The sun's relative elevation is measured with respect to the vertical and is known as its
17 "zenith angle." This angle varies hourly, seasonally, and with latitude. Daily and seasonal
18 changes in solar zenith angle result in the largest changes in the magnitude of solar radiation
19 flux, with higher zenith angles corresponding to lower solar flux. The largest natural fluxes
20 occur in the tropical regions, where solar noon occurs at a zenith angle at or near 0°. Seasonal
21 variation in solar flux ranges from small changes at the equator to very large changes at high
22 latitudes. Daily variations in solar flux, from sunrise to sunset, show added wavelength
23 dependence as a function of zenith angle, because transmission of some wavelengths are
24 sensitive to atmospheric pathlength due to scattering and absorption processes. These processes
25 will be discussed further below.
26
27 10.2.1.3 Atmospheric Radiative Interactions with Solar UV Radiation
28 Radiative Interactions in the Stratosphere
29 The stratosphere contains 90% or more of the total column density of O3, the principle gas
30 phase absorber of UV-B. Ozone interacts with UV radiation by scattering the photon or
31 absorbing and transforming its energy. Upon absorbing a UV photon, O3 may photodissociate,
32 or become electronically- and vibrationally-excited.
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1 Photoabsorption by O3 occurs with very high efficiency. After electronically-excited O3
2 (O3*) is formed, it will either dissociate into ground-state oxygen, O2, and an electronically
3 excited oxygen radical, O(1D), or a collision between the O3 molecule with another gas molecule
4 (M) may occur (See Reactions 1 and 2). Intermolecular collisions degrade the excess electronic
5 energy of the O3 molecule by transferring it to other molecules as vibrational, rotational, and/or
6 translational energies, which warm the atmosphere. The electronically excited oxygen radical
7 can then react with H2O to form two hydroxyl radicals, OH (Reaction 3). See Chapter 2 for
8 further discussion of odd oxygen and HOX photochemistry.
O3 + hv -> O3* -> O^D) + O2 (10-1)
-»O3*+M->O3 + M* (10-2)
9
10
O(1D) + H2O -» 2OH (10-3)
11
12
13 Either of these photochemical processes transforms the energy of the UV photon into a form
14 lacking the potential for human health or ecosystems damage.
15 The WMO/UNEP (2002) scientific assessment reported that global average total column
16 O3 had declined by 3% from pre-1980 levels, due to the presence of anthropogenic ozone-
17 depleting substances in the atmosphere. Ozone depletion has a strong latitude and seasonal
18 dependence. The seasonality of total O3 changes differ between the Northern and Southern
19 Hemispheres. In the northern midlatitudes, total column O3 declined by -4% during the
20 winter/spring seasons and by approximately half that amount in the summer/fall of the
21 1997-2001 time period, relative to pre-1980 total column O3 levels. In southern midlatitudes,
22 total column O3 declined -6% during all seasons.
23 The concentration of O3 in a vertical column extending from the earth's surface is
24 expressed in Dob son Units (DU), corresponding to the column height in hundredths of a
25 millimeter of O3 at standard temperature and pressure (273 K and 1 atmosphere) (Wayne, 2000).
26 One DU = 2.587 x 1015 molecules of O3/cm2. The total O3 column effectively prevents any
27 UV-C from reaching the surface, reduces the penetration of UV-B to the surface but does little to
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1 attenuate the intensity of UV-A except at the shorter wavelengths close to the cut off for UV-B.
2 Cutchis (1974) calculated that with overhead sun a 10% decrease in the O3 column would lead to
3 20, 250, and 500% increases in flux at 305, 290, and 287 nm, respectively, values that have been
4 supported by ground observations in Toronto, ON (49° N; Kerr and McElroy, 1993). Rapid
5 changes of this magnitude appear to happen naturally. As seen in data collected by the Total
6 Ozone Mapping Satellite (TOMS) (Figure 10-2), the total O3 column undergoes wide natural
7 variation on short timescales (Cockell, 2001).
Ozone Column (1990-1992)
600
c
o
(/)
£1
O
a
/ / / / / /
Date
Figure 10-2. Ozone column abundances from the years 1990 to 1992 for 0, 40, and 80° N
as well as 80° S. The data for 80° S are incomplete, but the graph shows the
effects of the Antarctic O3 hole on total column abundances at this latitude.
The data for the Northern Hemisphere illustrate the natural variations in the
O3 column over time. The data are taken from the TOMS (Total Ozone
Monitoring Satellite) data set (1979 to 1993).
Source: Cockell (2001).
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1 Nacreous and polar stratospheric clouds scatter radiation and aerosols, such as those
2 injected into the stratosphere by explosive volcanic eruptions, both absorb and scatter radiation.
3 Relative to the troposphere, the stratosphere is low in atmospheric pressure. Stratospheric clouds
4 and aerosols are also more dispersed than in the troposphere. Consequently, UV radiation can
5 traverse the stratosphere with a substantially lower probability of encountering a gas molecule or
6 cloud or aerosol particle than it would in the troposphere. In the radiative transfer literature, the
7 stratosphere is described as a "single scattering" regime for UV radiation, and UV that has
8 penetrated the stratosphere is referred to as "direct beam UV." The troposphere, due to its high
9 gas and particle concentrations is referred to as a "multiple scattering" regime.
10
11 Radiative Interactions in the Troposphere: Solar Irradiance vs. Actinic Flux
12 The troposphere contains < 10% of the total column O3 but -78% of the total atmospheric
13 mass including clouds, gas- and particle-phase radiation scatterers and absorbers, making it a
14 "multiple scattering" regime for UV radiation. These scattering processes increase the mean-
15 free path a photon must travel before reaching the surface, transforming the direct beam UV
16 solar irradiance that has penetrated the stratosphere into diffuse, or actinic, UV (Briihl and
17 Crutzen, 1989).
18 Atmospheric scattering processes, such as Rayleigh and Mie scattering, are particle-size
19 dependent. In Rayleigh scattering, gas molecules that are smaller than the wavelength of the
20 incident photon isotropically deflect incoming photons. With Mie scattering, aerosol and cloud
21 droplets scatter incoming radiation with forward- and back-scattering tendencies. Actinic flux,
22 especially at the earth's surface, is directly proportional to surface albedo (Wendisch and Mayer,
23 2003). Surface albedo is very strongly wavelength dependent. For example, fresh and wet snow
24 reflects 60 to 90% of violet light, while soil and grass surfaces reflect < 5% of incident violet
25 light (Xenopoulos and Schindler, 2001). Wendisch and Mayer (2003) found, in their in situ
26 measurement and modeling study of the vertical distribution of solar irradiance, that surface
27 albedos must be measured in order to accurately simulate solar flux, due to possibly large
28 variations in albedo within a given surface type. Snow cover, even many kilometers from
29 measurement sites is known to increase detected UV irradiances. Complicated interactions
30 result when radiation is scattered by snow (or other bright surfaces) and backscattered or
31 absorbed by atmospheric particles and clouds in the same vicinity (WMO/UNEP, 2002).
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1 Variation in Solar Flux with Altitude
2 Solar flux increases with altitude above sea level, due to the decreased presence of clouds
3 and declining concentrations of scattering and absorbing atmospheric pollutants. Rayleigh
4 scattering, discussed below, also lessens with decreasing atmospheric pressure. A number of
5 measurements of UV radiation have been taken as a function of altitude and are reviewed by
6 Xenopoulos and Schindler (2001). Increases in flux as a function of altitude are given as percent
7 irradiance enhancement per 1000 m relative to sea level. The effect can range from 9 to 24%
8 /1000 m as function of the altitude at which the measurement was taken (Xenopoulos and
9 Schindler, 2001). The effect corresponds to the relative pathlength traveled by the solar photon:
10 flux is strongest when the photon is not impeded by atmospheric scattering or absorbing agents.
11 Similarly, this effect is seen as a function of solar zenith angle, i.e., flux is at its maximum when
12 the atmospheric depth through which the photon must pass is at its shallowest.
13
14 Clouds
15 In principle, clouds have the largest influence on surface level UV irradiance, but their
16 effects are difficult to quantify. The depth and composition of a cloud determines, in part, the
17 amount and wavelengths of radiation that it will scatter or absorb. Geometry is an especially
18 important factor, as scattered or broken clouds may enhance, rather than reduce, surface flux.
19 For example, if the sun is not fully blocked by a cloud, the reduction in irradiance may be small
20 (WMO/UNEP, 2002). Quantifying the effect of clouds on surface UV flux, therefore, requires
21 information on cloud composition, geometry, altitude, and the position of the sun relative to the
22 cloud and the underlying surface, as a function of time. Provided that all of this detailed
23 information is available, a three dimensional model is then required to calculate surface-level
24 reductions or enhancements in UV flux.
25
26 Particulate Matter
27 On a zonally averaged basis, PM does not contribute significantly to lower tropospheric
28 absorption of UV radiation. However, in urban areas or other areas subject to high smog levels
29 (areas of significant biomass combustion), PM may be the most important determinant of
30 ground-level erythemal UV flux, second only to cloud cover (U.S. Environmental Protection
31 Agency, 2004; WMO/UNEP, 2003). Model-to-measurement comparisons of ground-level flux
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1 for Greece and Toronto have shown 20 and 5-10% reductions, respectively (McKenzie et al.,
2 2001). Increases over the past 20 to 30 years in combustion-associated PM and black carbon
3 may account for the inability to detect a surface trend in UV-B radiation caused by a known
4 decrease in stratospheric O3 over the Northern Hemisphere (Barnard et al., 2003).
5
6 Gases
1 In the upper troposphere, the UV-absorbing gases O3 and, of lesser importance,
8 formaldehyde (CH2O) and SO2 are vented or diffuse from the surface. Stratospheric intrusions
9 extrude O3-rich air into the troposphere where it mixes, increasing regional background O3 levels
10 (see Chapter 3). Tropospheric O3 data are typically expressed on a concentration basis, e.g.,
11 parts per billion by volume (ppbv), where 1 ppbv tropospheric O3 = 0.65 DU (IPCC, 2001). The
12 mean values of O3 in the free troposphere reported in the literature range from -50 to -80 ppbv,
13 with higher values occurring at the tropopause. For example, a series of ozonesonde soundings
14 over France during the period from 1976 to 1995 showed an O3 increase from 48.9 ppbv in the
15 2.5 to 3.5 km layer to 56.5 ppbv in the 6.5 to 7.5 km layer, although the data revealed no
16 significant increasing trend over time (Ancellet and Beekmann, 1997).
17 Close to the surface in polluted urban settings, photochemistry produces a diurnal rise and
18 fall in O3 and PM concentrations. Temperature inversions tend to prevent the upward mixing
19 and dilution of ground-level O3, while trapping primary and secondary PM. A recent study of
20 the concurrence of O3 and PM is provided by Koloutsou-Vakakis et al. (2001). No measurement
21 technique is currently available that can distinguish between absorption of incident UV radiation
22 by O3 versus absorption by PM.
23 UV absorption by gases becomes important under aerosol- and cloud-free conditions.
24 Figure 10-3 shows a calculation by Krotkov et al. (1998) of the sensitivity, as a function of
25 wavelength, of ground-level UV flux to a 1-DU decrease in total column O3 under cloud- and
26 aerosol-free conditions. A 1991 to 1992 study in Chicago in which ambient O3, broadband UV
27 irradiance and total sunlight were monitored (Frederick et al., 1993), produced a statistically
28 significant negative correlation between the UV irradiance and ambient O3 when the atmosphere
29 was relatively free of clouds and haze. Although they estimated that a 10-ppbv reduction in O3
30 was associated with a 1.3% increase in erythemally-weighted UV-B, they cautioned that this
31 figure had a comparatively large uncertainty (±1.2%).
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290
300 310 320
Wavelength (nm)
330
Figure 10-3. The sensitivity of ground-level UV flux to a 1 DU change in total column O3,
under clear sky conditions, as a function of solar zenith angle (SZA).
Source: Krotkov et al. (1998).
1
2
3
4
5
6
7
8
9
10
11
12
10.2.1.4 Modeling Surface UV-B Irradiance
The WMO/UNEP (2003) stated that, in principle, if the spatial distribution of all UV
absorbers and scatterers were fully known, the wavelength and angular distribution of the UV
irradiance at the earth's surface could be determined with model calculations. However, the
poorly known and complicated distribution of the primary components (clouds, particles, O3,
and surface albedo) makes detailed predictions extremely difficult.
Each of the factors described above must be integrated with human behaviors that also
influence potential exposure to UV-B radiation in order to calculate the human health risks
associated with small changes in total column O3 density. These human factors and the related
health consequences are described in detail below.
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1 10.2.2 Factors Governing Human Exposure to Ultraviolet Radiation
2 Multiple, complex factors determine the solar flux of UV radiation at the ground level, as
3 discussed in Section 10.2.1. Figure 10-4 illustrates some of these geophysical/atmospheric
4 factors including stratospheric and tropospheric O3, clouds, aerosols, and Rayleigh scattering.
5 An estimate of the impact of changes in surface-level O3 on UV-related human health effects
6 requires information about human physical, behavioral, and demographic factors associated with
7 exposure to UV radiation. In this section, the various factors that govern the degree of human
8 exposure to UV radiation are examined.
9
V
Backscattered
Radiation
Incident Solar UV Radiation
Stratospheric 03
Figure 10-4. Complexity of factors that determine human exposure to UV radiation.
In addition to the geophysical/atmospheric factors (e.g., stratospheric and
tropospheric O3, clouds, aerosols, and Rayleigh scattering) that affect the
solar flux of UV radiation at surface level, there are human physical,
behavioral and demographic factors that influence human exposure to
UV radiation.
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1 10.2.2.1 Outdoor Activities
2 Exposure to solar UV radiation is related to one predominating factor - time spent in
3 outdoor microenvironments during daylight hours. A large U.S. study was conducted using the
4 EPA National Human Activity Pattern Survey (NHAPS) to assess UV radiation dose in
5 Americans (Godar, 2001; Godar et al., 2001, 2003). The EPA NHAPS recorded the activity
6 profiles of 9,386 Americans (age 0 to 60+ years) over a 24-month period to assess their exposure
7 to various environmental pollutants, plus UV radiation. This study indicated that there was a
8 strong seasonal preference for outdoor activities, with people spending the most time outdoors
9 during the summer, followed by spring, fall, and, lastly, winter (Godar et al., 2001). Because the
10 solar erythemal (i.e., skin reddening) UV radiation dose is also highest during the summer, the
11 estimated UV radiation dose of Americans was more than ten-fold greater in the summer
12 compared to the winter season (Godar et al., 2001). Note that the error associated with
13 estimating UV radiation dose from exposure surveys and one EPA UV-monitoring site located at
14 each quadrant of the U.S. will be high.
15 Vacationing at the beach during the summer was associated with higher UV radiation
16 exposures (Godar et al., 2001; Thieden et al., 2001). Even after accounting for sunscreen use at
17 the beach, the erythemal UV radiation doses were more than 40% higher during a three-week
18 beach vacation compared to a three-week stay at home (Godar et al., 2001). Danish children and
19 adolescents were found to receive > 50% of their annual UV radiation dose while vacationing at
20 European beaches (Thieden et al., 2004a). The high UV radiation dose received at beaches is
21 due to increased time spent outdoors during daylight hours and increased risk behavior, namely
22 sunbathing. Sunbathing also was associated with increased annual UV radiation dose in the
23 Canadian National Survey on Sun Exposure and Protective Behaviours (Shoveller et al., 1998).
24 Among the 3,449 adults (age 25+ years) who completed the telephone household survey, 21%
25 stated that they spent time actively sunbathing. In a Danish study with 164 participants, all
26 children (age 1 to 12 years) and teenagers (age 13 to 19 years) and 94% of adults (age 20 to
27 76 years) had days with risk behavior (Thieden et al., 2004b). Teenagers had the highest annual
28 UV radiation doses, as compared to children and adults, a finding likely attributable to their
29 having the highest number of risk-behavior days. Among teenagers, 76% of their UV radiation
30 dose during the measurement period was received on risk-behavior days, as determined using
31 personal electronic UV dosimeters and exposure diaries (Thieden et al., 2004b). In addition to
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1 vacationing and sunbathing, participation in outdoor sports (e.g., basketball, soccer, golfing,
2 swimming, cycling) also significantly increased UV radiation exposure (Moehrle, 2001; Moehrle
3 et al., 2000; Moise et al., 1999; Thieden et al., 2004a,b).
4
5 10.2.2.2 Occupation
6 Of the various factors that affect human exposure to UV radiation, occupation is also
7 important. Approximately 5% of the American workforce perform work in outdoor
8 microenvironments, as determined by the EPA NHAPS (Godar et al., 2001). American indoor
9 workers, on average, spend -10% of their day outdoors and are exposed to -30% of the total
10 ground-level UV flux, as measured by the EPA UV-monitoring program, during this time period
11 (Godar et al., 2001). In comparison to indoor or home workers, outdoor workers are exposed to
12 much higher levels of UV radiation (Kimlin et al., 1998a; Thieden et al., 2004a), frequently at
13 levels that are above current exposure limits set by the International Commission on
14 Non-Ionizing Radiation Protection (ICNIRP, 2004). For example, Thieden et al. (2004a)
15 observed that the annual UV radiation dose, estimated using personal electronic UV dosimeters
16 and exposure diaries, -70% higher for gardeners than indoor workers. The gardeners received
17 the majority (55%) of their UV radiation dose on working days (Thieden et al., 2004a). Another
18 study found that outdoor workers received three to four times the annual UV radiation exposure
19 of indoor workers (Diffey, 1990). At-risk working populations include farmers (Airey et al.,
20 1997; Schenker et al., 2002), fisherman (Rosenthal et al., 1988), landscapers (Rosenthal et al.,
21 1988), building and construction workers (Gies and Wright, 2003), physical education teachers
22 (Vishvakarman et al., 2001), mail delivery personnel (Vishvakarman et al., 2001), and various
23 other workers who spend the majority of their day in outdoor microenvironments during peak
24 UV radiation hours.
25
26 10.2.2.3 Age
27 Age-related differences in UV radiation dose was examined in a U.S. study using the EPA
28 NHAPS (Godar, 2001; Godar et al., 2001). The average UV radiation dose among American
29 children (age < 12 years) was estimated to be similar to that of adults (age 20+ years), but -20%
30 higher than that of adolescents (age 13 to 19 years) (Godar, 2001). During the summer, higher
31 UV radiation doses were observed in children < 5 years of age and men 40+ years of age (Goder,
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1 2001). In a large Canadian survey, 89% of children (age < 12 years) had 30 minutes or more of
2 daily UV exposure, compared to 51% for both adults (age 25+ years) and youth (age 15 to
3 24 years) (Lovato et al., 1998a, 1998b; Shoveller et al., 1998). In an English study (Diffey et al.,
4 1996), UV radiation exposure was estimated in 180 children (age 9 to 10 years) and adolescents
5 (age 14 to 15 years) using personal film badges and exposure records. Once again, children were
6 found to have received higher UV radiation exposure compared to adolescents (Diffey et al.,
7 1996). However, in contrast to the results from the studies discussed above, a Danish study
8 observed that the annual UV radiation dose in teenagers (age 13 to 19 years) was -20% higher
9 compared to children (age 1 to 12 years) (Thieden et al., 2004b). This increase in UV radiation
10 dose in teenagers was attributed to their increased risk-behavior days. The time profiles of daily
11 UV radiation exposure among grade 8 students was assessed in an Australian study using a
12 routinely operating UV-Biometer and questionnaires (Moise et al., 1999). The results indicated
13 that up to 47% of the daily UV radiation dose fell within the time periods when students were
14 outdoors during school hours, sitting under shaded structures during lunch breaks and
15 participating in routine outdoors or sports activities (Moise et al., 1999).
16 Two studies examined lifetime UV radiation exposure among persons in the U.S. (Godar
17 et al., 2001) and Denmark (Thieden et al., 2004b). Both studies observed that while there are
18 slight differences in UV radiation dose by age, generally people receive fairly consistent UV
19 doses at different age intervals throughout their lives.
20
21 10.2.2.4 Gender
22 Studies have indicated that females generally spend less time outdoors and, consequently,
23 have lower UV radiation exposure compared to males (Gies et al., 1998; Godar et al., 2001;
24 Shoveller et al., 1998). The U.S. study by Godar et al. (2001) observed that even though both
25 males and females had somewhat consistent erythemal UV radiation doses throughout their
26 lives, males received consistently higher UV doses compared to females at all age groups.
27 Among all Americans, the lowest exposure to UV radiation was received in females during their
28 child-raising years (age 22 to 40 years) (Godar et al., 2001). The highest exposure was observed
29 in males aged 41 to 59 years in the U.S. study (Godar et al., 2001), while a similar Canadian
30 survey found that younger adult males had the greatest exposures to UV radiation (Shoveller
31 etal., 1998).
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1 10.2.2.5 Geography
2 In the U.S. study by Godar et al. (2001), erythemal UV radiation doses were examined in
3 persons living in northern and southern regions. Northerners and southerners were found to
4 spend an equal amount of time outdoors; however, the higher solar flux at lower latitudes
5 significantly increased the annual UV radiation dose for southerners (Godar et al., 2001). The
6 annual UV radiation doses in southerners were 25 and 40% higher in females and males,
7 respectively, compared to northerners (Godar et al., 2001). Other studies also have shown that
8 altitude and latitude influence personal exposure to UV radiation (Kimlin et al., 1998b; Rigel
9 etal., 1999).
10
11 10.2.2.6 Protective Behavior
12 Protective behaviors such as using sunscreen (e.g., Nole and Johnson, 2004), wearing
13 protective clothing (e.g., Rosenthal et al., 1988; Sarkar, 2004; Wong et al., 1996), and spending
14 time in microenvironments with shaded surfaces (Moise et al., 1999; Parisi et al., 1999) have
15 been shown to reduce exposure to UV radiation. In one study, the use of sunscreen was
16 associated with extended intentional UV radiation exposure (Autier et al., 1999); however, a
17 follow-up study indicated that sunscreen use increased duration of exposures to doses of UV
18 radiation that were below the threshold level for erythema (Autier et al., 2000).
19 In a national study of U.S. youths aged 11 to 18 years, the most prevalent protective
20 behavior was sunscreen use (39.2%) followed by use of a baseball hat (4.5%) (Davis et al.,
21 2002). There were significant differences in the use of sunscreen by age group and gender, with
22 the younger age group (age 11 to 13 years) and girls having greater likelihood (47.4 and 48.4%,
23 respectively) of using sunscreen (Davis et al., 2002). The Canadian National Survey on Sun
24 Exposure and Protective Behaviours observed that less than half of the adults (age 25+ years,
25 n = 3,449) surveyed took adequate protective actions (Shoveller et al., 1998). Once again,
26 children (age < 12 years, n = 1,051) were most protected from exposure to UV radiation, with
27 76% using sunscreen and 36% avoiding the sun, as reported by their parents (Lovato et al.,
28 1998a). However, the protection level was still not adequate, as indicated by the high 45% rate
29 of erythema in children. Among Canadian youth (age 15 to 24 years, n = 574), protective
30 actions from UV radiation exposure included wearing a hat (38%) and seeking shade and
31 avoiding the sun between the peak hours of 11:00 a.m. to 4:00 p.m. (26%) (Lovato et al., 1998b).
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1 The lowest prevalence of protective behavior among the youth was likely responsible for the
2 highest proportion of erythema (68%) experienced in this age group. A Danish study observed
3 that both children and teenagers applied sunscreen on more days than adults, but teenagers had
4 the most days with erythema, due to their increased risk behavior (Thieden et al., 2004b).
5 A survey in Switzerland of 1,285 individuals, including children and parents, indicated that
6 sunscreen use was the protective action most commonly used, but only at the beach and not in
7 routine daily exposure (Berret et al., 2002). Protective clothing and avoiding the sun were not
8 highly used among these individuals.
9
10 10.2.2.7 Summary of Factors that Affect Human Exposures to Ultraviolet Radiation
11 The factors that potentially influence UV radiation doses were discussed in the previous
12 sections and include choice of leisure activities, occupation, age, gender, geography, and
13 protective behavior. Results from the various studies indicate that specific populations may be at
14 risk for higher exposures to UV radiation. Of particular concern are the following potentially
15 susceptible populations:
16 • Individuals who engage in high-risk behavior, viz., sunbathing;
17 • Individuals who participate in outdoor sports and activities, including professional
athletes;
18 • Individuals who work in outdoor microenvironments with inadequate shade, e.g.,
farmers, fishermen, gardeners, landscapers, building and construction workers;
19 • Young children; and
20 • Individuals living in geographic areas with higher solar flux (i.e., lower latitudes
[e.g., Honolulu, HI] and higher altitudes [e.g., Denver, CO]).
21
22 10.2.3 Factors Governing Human Health Effects due to Ultraviolet Radiation
23 Ultraviolet radiation occupies a specific region of the electromagnetic spectrum of
24 wavelengths and can be further subdivided into three parts, UV-A (320 to 400 nm), UV-B
25 (280 to 320 nm), and UV-C (200 to 280 nm). Most of the health risks associated with UV
26 radiation exposure are wavelength dependent. Wavelengths < 180 nm are of little practical
27 biological significance as they are readily absorbed in the air (ICNIRP, 2004). Until 1980,
28 it was generally thought that wavelengths < 315 nm were responsible for the most significant
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1 adverse UV radiation health effects; however, recent studies have found that the longer
2 wavelengths in the UV-A range also might produce adverse responses at substantially higher
3 doses (ICNIRP, 2004).
4 Action spectra of a given biological response to UV radiation across its spectral range are
5 used to estimate exposure by weighting individual wavelength intensities according to the
6 associated response. The overall effectiveness of the incident flux at inducing the biological
7 response of interest is computed by means of the relationship
8
effective if radiance = )I^E^a"X (10-4)
A
9
10 where 1A and EA are, respectively, the irradiance and its relative effectiveness at wavelength A.
1 1 UV-A and UV-B radiations differ in their abilities to initiate DNA damage, cell signaling
12 pathways, and immune alterations. In this section, the various adverse health effects associated
13 with acute and chronic UV radiation exposure will be discussed.
14
15 10.2.3.1 Erythema
1 6 Association Between Ultraviolet Radiation Exposure and Erythema
17 The most conspicuous and well-recognized acute response to UV radiation is erythema, or
18 the reddening of the skin, which is likely caused by direct damage to DNA by UV-B and UV-A
19 radiation (Matsumura and Ananthaswamy, 2004). Indirect oxidative damage also might occur at
20 longer wavelengths (Matsumura and Ananthaswamy, 2004). Skin type appears to play a large
21 role in the sensitivity to UV radiation -induced erythema. The Fitzpatrick classifications for skin
22 types are: (1) skin type I - individuals with extremely sensitive skin that sunburns easily and
23 severely, and is not likely to tan (e.g., very fair skin, blue eyes, freckles); (2) skin type
24 II - individuals with very sensitive skin that usually sunburns easily and severely, and tans
25 minimally (e.g., fair skin, red or blond hair, blue, hazel or brown eyes); (3) skin type
26 III - individuals with sensitive skin that sunburns moderately and tans slowly (e.g., white skin,
27 dark hair); (4) skin type IV - individuals with moderately sensitive skin that sunburns minimally
28 and usually tans well (e.g., white or light brown skin, dark hair, dark eyes); (5) skin type
29 V - individuals with minimally sensitive skin that rarely sunburns and tans deeply (e.g., brown
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1 skin); and (6) skin type VI - individuals with nonsensitive skin that never sunburns and tans
2 profusely (e.g., dark skin). A study by Harrison and Young (2002) found that the perceptible
3 minimal erythemal dose was approximately two-fold greater among individuals with skin type
4 IV compared to skin type I, although there was considerable overlap in the minimal erythemal
5 dose among the four skin types. Waterston et al. (2004) further observed that within an
6 individual, erythemal response differed by body site (e.g., abdomen, chest, front upper arm, back
7 of thigh). These differences were likely attributable to body site-specific variations in melanin
8 pigmentation.
9 Kollias et al. (2001) investigated the change in erythemal response following a previous
10 exposure to UV radiation. Body sites that received a second exposure to UV radiation always
11 showed a reduced erythemal response compared to body sites with a single exposure, especially
12 when the first exposure was at levels greater than the minimal erythemal dose. The suppression
13 of erythema was more pronounced when the second exposure was given 48 hours after the first.
14 However, Kaidbey and Kligman (1981) found that multiple exposures to subthreshold doses of
15 UV radiation at 24-hour intervals were found to lower the minimal erythemal dose. Henrisken
16 et al. (2004) observed that the change in threshold depended on skin type. In 49 healthy
17 volunteers with skin types II, III, and IV, just perceptible erythema 24 hours post-exposure was
18 chosen as the minimal erythemal dose. After four days of repeated UV radiation, the minimal
19 erythemal dose was lowered by 40 to 50% in darker-skinned persons. However, among fair-
20 skinned individuals, there was no change in the erythemal threshold dose with repeated exposure
21 to UV radiation.
22 A reference erythema action spectrum was adopted by the Commission Internationale de
23 1'Eclairage (International Commission on Illumination, CIE) in 1987 (McKinlay and Diffey,
24 1987). The CIE erythema action spectrum indicates that UV-B radiation is orders of magnitude
25 more effective per unit dose than UV-A radiation. However, a follow-up study by Diffey (1994)
26 that compared the observed and predicted minimal erythemal doses found that the erythemal
27 sensitivity of skin at longer UV wavelengths (> 350 nm) was greater than predicted from the CIE
28 reference action spectrum.
29
30
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1 Risk of Erythema from Changes in Tropospheric Ozone Levels
2 There is no literature on the risk of erythema associated with changes in tropospheric or
3 ground-level O3 levels. However, one study conducted a risk assessment of the effects of
4 stratospheric O3 depletion on the risk of erythema (Longstreth et al., 1998). Stratospheric O3
5 depletion will result in only a slight increase influence rate of UV-A as O3 only absorbs a very
6 little part of this UV spectrum. However, the ground-level UV-B flux will likely increase as O3
7 absorbs radiation in that wavelength range with high efficiency.
8 The risk analysis by Longstreth et al. (1998) concluded that erythema will not appreciably
9 increase with depletion of the O3 layer. This is due to the powerful adaptation of the skin to
10 different levels of UV radiation, as evidenced by its ability to cope with changes in UV radiation
11 by season (van der Leun and de Gruijl, 1993). Gradual exposure to increasing UV radiation
12 from the winter to summer leads to decreased sensitivity of the skin. In midlatitudes, the UV-B
13 radiation in the summer is ten-fold greater than in the winter. In contrast, the steady depletion of
14 the O3 layer has been estimated to likely lead to an -20% increase in UV-B in 10 years
15 (Longstreth et al., 1998). Such a comparatively small increase in UV radiation throughout the
16 years, therefore, would not seem likely to significantly increase the risk of erythema. Given that
17 the stratospheric O3 depletion is estimated as not being likely to affect the risk of erythema, one
18 can conclude that changes in ground-level O3 (which only constitutes no more than 10% of total
19 atmospheric O3) is not likely to result in increased risk.
20
21 10.2.3.2 Skin Cancer
22 According to the Skin Cancer Foundation, one in six Americans will develop skin cancer
23 during their lifetime (Gloster and Brodland, 1996). The three main forms of skin cancer include
24 basal cell carcinoma and squamous cell carcinoma, which are both nonmelanoma skin cancers,
25 and malignant melanoma. Nonmelanoma skin cancers constitute more than one-third of all
26 cancers in the U.S. and -90% of all skin cancers, with basal cell carcinoma being approximately
27 four times as common as squamous cell carcinoma (Diepgen and Mahler, 2002; ICNTRP, 2004).
28 The incidence of malignant melanoma (3 to 4% of all cancers) is much lower compared to
29 nonmelanoma skin cancers, but melanoma has great metastatic potential and accounts for the
30 majority of skin cancer deaths (Jemal et al., 2004).
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1 Exposure to UV radiation is considered to be a major risk factor for all three forms of skin
2 cancer (Gloster and Brodland,1996; Diepgen and Mahler, 2002; IARC, 1992). Ultraviolet
3 radiation is especially effective in inducing genetic mutations and acts as both a tumor initiator
4 and promoter. Keratinocytes have evolved DNA repair mechanisms that correct the damage
5 induced by UV, but mutations do occur, leading to skin cancers that are appearing with
6 increasing frequency (Hildesheim and Fornace, 2004). The relationship between skin cancer and
7 chronic exposure to UV radiation is further explored below, followed by discussion of the
8 influence of O3 on the incidence of skin cancer.
9
10 10.2.3.3 Ultraviolet Radiation Exposure and the Incidence of Nonmelanoma
11 Skin Cancers
12 The incidence of all three types of cancers has been shown to rise with increasing UV
13 radiation concentrations across the U.S. (de Gruijl, 1999); however, the most convincing
14 evidence for a causal relationship exists between UV radiation and squamous cell carcinoma.
15 Squamous cell carcinoma occurs almost exclusively on skin that is regularly exposed to the sun,
16 such as the face, neck, arms, and hands. The incidence is higher among whites in areas of lower
17 latitudes, where solar flux is greater (Kricker et al., 1994). The risk of squamous cell carcinoma
18 was shown to increase with life-long accumulated exposure to UV radiation in one cross-
19 sectional study (Vitasa et al., 1990), but was found to be associated only with the exposure ten
20 years prior to diagnosis in a case-control study (Gallagher et al., 1995a). One of the major
21 concerns with both types of studies is the potential for recall bias in reporting their UV radiation
22 exposure as individuals are already aware of their disease status.
23 Ultraviolet radiation also has been linked to basal cell carcinoma. Basal cell carcinoma is
24 common on the face and neck (80-90%) but rarely occurs on the back of the hands (de Gruijl,
25 1999). While cumulative UV radiation exposure was not associated with risk of basal cell
26 carcinoma (Vitasa et al., 1990), increased risk was observed in individuals with increased
27 recreational UV radiation exposure in adolescence and childhood (age < 19 years) and
28 individuals with a history of severe erythema in childhood (Gallagher et al., 1995b). Once again,
29 consideration must be given to potential recall bias in assessing these results. Thus, there is
30 suggestive evidence that UV radiation also plays a role in the development of basal cell
31 carcinoma, but the etiologic mechanisms for squamous cell carcinoma and basal cell carcinoma
32 likely differ. In an Australian study conducted in a subtropical community, having fair skin,
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1 a history of repeated sunburns, and nonmalignant solar skin damage diagnosed by dermatologists
2 were strongly associated with both types of nonmelanoma skin cancer (Green et al., 1996).
3 Outdoor occupation was not associated with nonmelanoma skin cancer, which was likely due to
4 significant self-selection. Individuals with fair or medium complexions and a tendency to
5 sunburn, though they accounted for more than 80% of the community study sample, were
6 systematically underrepresented among outdoor workers (Green et al., 1996). This selection bias
7 may partly explain the lack of consistent quantitative evidence of a causal link between UV
8 radiation and skin cancer in humans.
9 De Gruijl et al. (1993) assessed the action spectrum for nonmelanoma skin cancers using
10 hairless albino mice. Human data are not available to examine the wavelength dependence of
11 the carcinogenicity of UV radiation. After adjusting for species differences, the Skin Cancer
12 Utrecht-Philadelphia action spectrum indicated the highest effectiveness in the UV-B range with
13 a maximum at 293 nm, which dropped to 10~4 of this maximum at the UV-A range above
14 340 nm (de Gruijl et al., 1993). The mutations commonly present in thep53 tumor suppressor
15 gene in individuals with squamous cell carcinoma and basal cell carcinoma are called the
16 "signature" mutations of UV-B radiation (de Gruijl, 2002). UV-B radiation is highly mutagenic
17 due to the fact that DNA is a chromophore for UV-B, but not for UV-A, radiation (Ichihashi
18 et al., 2003). However, other studies have found that UV-A radiation, in addition to UV-B
19 radiation, can induce DNA damage (Persson et al, 2002; Ruenger et al., 2000). DNA damage by
20 UV-A is mediated by reactive oxygen species, thus is indistinguishable from damage caused by
21 other agents that generate reactive oxygen species (de Gruijl, 2002). Epidemiologic evidence of
22 a carcinogenic effect of UV-A was found in a study of psoriasis patients receiving oral psoralen
23 and UV-A radiation treatment (Stern et al., 1998). High-dose exposure to oral psoralen and
24 UV-A radiation was associated with a persistent, dose-related increase in the risk of squamous
25 cell cancer, but with much less effect on the risk of basal cell cancer. Therefore, although UV-B
26 radiation has long been considered the main culprit for nonmelanoma skin cancer, limited
27 evidence suggest that UV-A radiation may also play a role.
28 Susceptible populations for nonmelanoma skin cancers include individuals with reduced
29 capacity for nucleotide excision repair, the primary repair mechanism for UV radiation-induced
30 DNA lesions (Ichihashi et al., 2003). At particular risk are individuals with xeroderma
31 pigmentosum, as they have defective nucleotide excision repair in all tissues (Kraemer, 1997;
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1 Sarasin, 1999). Skin type also largely affects susceptibility to skin cancer. Of the six skin
2 phenotypes, the most sensitive individuals are those with skin types I and II, who have a fair
3 complexion, blue or green eyes, and red or blond hair (Diepgen and Mahler, 2002). These
4 individuals tend to sunburn easily, tan poorly, and freckle with sun exposure. A history of
5 repeated sunburns also appears to increase the risk of both cancers, while sunburns during
6 childhood are more associated with increased basal cell carcinoma (Gallagher et al., 1995b;
7 Green etal., 1996).
8
9 Ultraviolet Radiation and the Incidence of Cutaneous Malignant Melanoma
10 From 1973 to 1994, the incidence rate of melanoma increased 120.5% along with an
11 increased mortality rate of 38.9% among whites in the U.S. (Hall et al., 1999). The ICNIRP
12 (2004) states that during the past 40 years or so, each decade has seen a two-fold increase in the
13 incidence of malignant melanoma in white populations, with increased incidence observed more
14 prominently in individuals living in lower latitudes. Cutaneous malignant melanoma has
15 mutifactorial etiology, with environmental, genetic, and host factors (Lens and Dawes, 2004).
16 While the major environmental factor of melanoma has been identified as UV radiation
17 exposure, the risk of melanoma appears to depend on the interaction between the nature of the
18 exposure and skin type (Lens and Dawes, 2004).
19 Fears et al. (2002) examined the association between invasive cutaneous melanoma and
20 UV radiation in non-Hispanic whites using a case-control study design. Lifetime residential
21 history was coupled with mid-range UV-B radiation flux measurements to reduce exposure
22 misclassification and recall bias. A 10% increase in the average annual UV-B flux was
23 significantly associated with a 19% increase in individual odds for melanoma in men and a
24 16% increase in women. Whiteman et al. (2001) conducted a systematic review of studies that
25 examined the association between childhood UV radiation exposure and risk of melanoma.
26 Researchers found that ecological studies assessing ambient sun exposure consistently reported
27 higher risks of melanoma among people who resided in an environment with high UV radiation
28 during their childhood (Whiteman et al., 2001). The lack of consistency among the case-control
29 studies was likely due to the varying methods used to assess UV radiation dose.
30 While the evidence is generally suggestive of a causal relationship between UV radiation
31 and malignant melanoma, possibly conflicting data has been observed. For example, the highest
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1 occurrence of malignant melanoma is on men's backs and women's legs, areas that do not have
2 prolonged exposure to the sun (Rivers, 2004). This indicates that malignant melanoma tends to
3 occur in sites of intermittent, intense sun exposure (trunk and legs), rather than in areas of
4 cumulative sun damage (head, neck, and arms), in contrast to nonmelanoma skin cancers
5 (Swetter, 2003).
6 The available data conflict with regard to the relative importance of UV-A versus UV-B in
7 inducing melanomas. UV-A has a much higher flux rate at the earth's surface, as it is not
8 absorbed by O3 and it is able to penetrate more deeply into the skin surface due to its longer
9 wavelength. However, UV-B, as mentioned earlier, is much more energetic and, therefore, more
10 effective in photochemically altering DNA. The individual roles of UV-B and UV-A in the
11 development of cutaneous malignant melanoma have been examined in several studies.
12 A case-control study of 571 patients and 913 matched controls observed an elevated odds ratio
13 for development of malignant melanoma in individuals who regularly used tanning beds, which
14 typically are UV-A sources (Westerdahl et al., 2000). In a study by Setlow et al. (1993),
15 an action spectrum using the tropical fish Xiphophorus indicated that UV-A range wavelengths
16 were especially important in malignant melanoma induction. However, an action spectrum
17 using the opossum Monodelphis domestica found that the potency of UV-A for melanoma
18 induction was extremely low compared to that of UV-B (Robinson et al., 2000). A recent study
19 by De Fabo et al. (2004) examined the differences in wavelength effectiveness using a
20 hepatocyte growth factor/scatter factor-transgenic mouse model. The epidermal tissue of these
21 transgenic mice behaves very similarly to the human epidermis in response to UV exposure.
22 Given the absence of a mammalian melanoma action spectrum, the standardized CIE erythema
23 action spectrum was used to deliver identical erythemally effective doses. Only UV-B radiation
24 was found to initiate mammalian cutaneous malignant melanoma. UV-A radiation, even at
25 doses considered physiologically relevant, were ineffective at inducing melanoma (De Fabo
26 et al., 2004). Overall, current evidence suggests that UV-B, and not UV-A, is the primary risk
27 factor for malignant melanoma (ICNIRP, 2004).
28 The populations susceptible for malignant melanoma are similar to those for nonmelanoma
29 skin cancers. Once again, individuals with xeroderma pigmentosum or a reduced capacity of
30 nucleotide excision repair are at increased risk (Tomescu et al., 2001; Wei et al., 2003).
31 Individuals with skin types I and II, or the fair-skin phenotype (blue or green eyes; blond or red
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1 hair; skin that freckles, sunburns easily, and does not tan), have increased susceptibility to
2 malignant melanoma (Evans et al., 1988; Swetter, 2003; Veier0d et al., 2003). However, the
3 incidence of melanoma was also positively associated with UV radiation in Hispanics and blacks
4 (Hu et al., 2004). Although the incidence of melanoma is much lower in Hispanics and blacks
5 compared to whites, melanomas in these populations are more likely to metastasize and have a
6 poorer prognosis (Black et al., 1987; Bellows et al., 2001). Among children, malignant
7 melanoma appears to have similar epidemiologic characteristics to the adult form of the disease
8 (Whiteman et al., 1997). Individuals with intermittent, intense sun exposure, particularly during
9 childhood, were found to have increased risk of melanoma (Whiteman et al., 2001), in contrast
10 to the association between cumulative exposure and increased risk of squamous cell carcinoma.
11 One study found that a personal history of nonmelanoma skin cancer or precancer, higher
12 socioeconomic status, and increased numbers of nevocytic nevi also were associated with
13 increased incidence of melanoma (Evans et al., 1988).
14
15 Effect of Changes in Tropospheric Ozone Levels on Skin Cancer Incidence
16 The current evidence strongly suggests a causal link between exposure to UV radiation and
17 the incidence of both nonmelanoma and melanoma skin cancer. Genetic factors, including skin
18 phenotype and ability to repair DNA, affect an individual's susceptibility to skin cancer.
19 Quantifying the relationship between UV radiation and skin cancer is complicated by the
20 uncertainties involved in the selection of an action spectrum and appropriate characterization of
21 dose (e.g., peak or cumulative levels of exposure, childhood or lifetime exposures). In addition,
22 there are multiple complexities in attempting to quantify the effect of tropospheric O3 levels on
23 UV-radiation exposure, as described in Section 9.11.2. There is an absence of published studies
24 that critically examine any increased incidence of skin cancer that may be attributable to
25 decreased tropospheric O3 exposures (which reflects the significant challenges in determining
26 ground-level O3-related changes in UV radiation exposure).
27 Several studies have examined the potential effect of stratospheric O3 depletion on the
28 incidence of skin cancer (de Gruijl, 1995; Longstreth et al., 1995; Madronich and de Gruijl,
29 1993; Slaper et al., 1996). Stratospheric O3 depletion is likely to increase the ground-level UV-B
30 flux, as O3 absorbs radiation in that wavelength range with high efficiency. Because UV-B
31 radiation is primarily implicated in the induction of skin cancer, especially among persons with
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1 skin phenotypes I and II, there is concern that the depletion of the O3 layer will result in
2 significantly increased incidence of skin cancers.
3 Estimation of the increased risk in melanoma associated with O3 depletion cannot be done
4 adequately due to the lack of a mammalian action spectrum for melanoma. In addition, the
5 complexity of the UV-related induction mechanism of melanoma adds an additional layer of
6 uncertainty to the calculations. The excess risk in nonmelanoma skin cancers associated with a
7 decrease in stratospheric O3 was estimated using the Skin Cancer Utrecht-Philadelphia action
8 spectrum based on hairless albino mice (Longstreth et al., 1995). Quantification of how much
9 more UV radiation would reach ground level with each percentage decrease in O3 required
10 several assumptions: (1) annual doses are an appropriate measure; (2) personal doses are
11 proportional to ambient doses; and, most notably, (3) each percentage decrease in O3 is
12 associated with a 1.2% increase in UV radiation. Next, the relationship between UV radiation
13 and nonmelanoma skin cancer incidence was determined. Each percent increase in annual UV
14 radiation dose was estimated to cause a 2.5% increase in squamous cell carcinoma and 1.4%
15 increase in basal cell carcinoma over a human lifetime. Incorporating all these factors,
16 Longstreth et al. (1995) calculated that a sustained 10% decrease in stratospheric O3
17 concentration would result in 250,000 additional nonmelanoma skin cancer cases per year.
18 Madronich and de Gruijl (1993) noted that the largest percent of O3-induced nonmelanoma skin
19 cancer increases would be at high latitudes, where baseline incidence of skin cancer is usually
20 small. Assuming a phaseout of primary O3-depleting substances by 1996, as established by the
21 Copenhagen Amendments in 1992, Slaper et al. (1996) estimated that the number of excess
22 nonmelanoma skin cancers in the U.S. caused by O3 depletion would exceed 33,000 per year
23 (or ~7 per 100,000) around the year 2050.
24 However, estimating the increase in nonmelanoma skin cancer incidence attributable to the
25 depletion of the stratospheric O3 layer is marred by uncertainty. The following statement by
26 Madronich and de Gruijl (1994) describes the uncertainty of estimating the effect of
27 stratospheric O3 depletion on the incidence of skin cancer:
z9 Extrapolating trends and effects of UV into the future is very hypothetical due to
30 uncertainties that arise from atmospheric chemistry, epidemiology, and related
31 disciplines. The values that we calculated are one plausible measure of the
32 magnitude of the 03-UV effects... .The timescales for atmospheric change and skin-
33 cancer development are still far from certain: 03 reductions are expected to continue
34 well into next century, and the time between UV exposure and development of skin
35 cancer is essentially unknown. . .
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1 Therefore, much caution is necessary when assessing and interpreting the quantitative results of
2 excess nonmelanoma skin cancer incidence due to stratospheric O3 depletion. Note that although
3 the effect of changes in ground-level or tropospheric O3 concentrations on skin cancer incidence
4 has not been assessed, it would be expected to be much less compared to the effect from the
5 depletion of the stratospheric O3 layer, given that tropospheric O3 makes up < 10% of the total
6 atmospheric O3.
7
8 10.2.3.4 Ocular Effects of Ultraviolet Radiation Exposure
9 Ultraviolet Radiation Exposure and Risk of Ocular Damage
10 Ocular damage from UV radiation exposures includes effects on the cornea, lens, iris, and
11 associated epithelial and conjunctival tissues. Absorption of UV radiation differs by
12 wavelength, with short wavelengths (< 300 nm) being almost completely absorbed by the
13 cornea, whereas longer wavelengths are transmitted through the cornea and absorbed by the lens
14 (McCarty and Taylor, 2002). The most common acute ocular effect of environmental UV
15 exposure is photokeratitis, also known as snowblindness, caused by absorption of short
16 wavelength UV radiation by the cornea. The action spectrum indicated that maximum
17 sensitivity of the human eye was found to occur at 270 nm (ICNIRP, 2004; Pitts, 1993). The
18 threshold for photokeratitis in humans varied from 4 to 14 mJ/cm2 for wavelengths 220 to
19 310 nm. For UV-A radiation, levels exceeding 10 J/cm2 were necessary for corneal injury.
20 Exposure to longer wavelengths has been shown to cause both transient and permanent
21 opacities of the lens, or cataracts. There is extensive toxicologic and epidemiologic evidence
22 supporting the causal association between UV radiation and cataracts (Hockwin et al., 1999;
23 McCarty and Taylor, 2002). Ultraviolet radiation-induced cataracts are hypothesized to be
24 caused by oxidative stress leading to increased reactive species in the lens, which then causes
25 damage to lens DNA and cross-linking of proteins. Exposure time to low-dose UV radiation was
26 found to strongly influence cataract formation (Ayala et al., 2000). An action spectrum
27 determined using young female rats indicated that the rat lens was most sensitive to 300 nm,
28 correcting for corneal transmittance (Merriam et al., 2000). Oriowo et al. (2001) examined the
29 action spectrum for cataract formation using whole cultured lens from pigs. As pigs lens are
30 similar in shape and size to the human lens, some inferences may be made. Results indicated
31 that the 270 to 315 nm waveband was most effective in producing UV-induced cataracts in vitro.
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1 However, the threshold values varied widely within that range, from 0.02 J/cm2 for 285 nm to
2 0.74 J/cm2 for 315 nm (Oriowo et al., 2001). At wavelengths > 325 nm, the threshold levels
3 were orders of magnitude larger, with a minimum threshold value of 18.7 J/cm2. A U.S. study of
4 838 watermen found that UV-B radiation was significantly associated with cortical, but not
5 nuclear, cataract formation (Taylor et al., 1988). No association was observed between cataracts
6 and UV-A radiation in this outdoor-working population.
7
8 Risk of Ocular Damage from Changes in Tropospheric Ozone Levels
9 Cataracts are the most common cause of blindness in the world. McCarty et al. (2000)
10 calculated that ocular UV radiation exposure accounted for 10% of the cortical cataracts in an
11 Australian cohort of 4,744 individuals from both urban and rural areas. A study by Javitt and
12 Taylor (1994-1995) found that the probability of cataract surgery in the U.S. increased by 3% for
13 each IE decrease in latitude. These results suggest that depletion of the stratospheric O3 layer
14 may increase UV radiation-induced cataract formation. After assuming a certain wavelength
15 dependency along with some additional assumptions, every 1% decrease in the stratospheric O3
16 layer was estimated to be associated with a 0.3 to 0.6% increase in cataracts (Longstreth et al.,
17 1995). Longstreth et al. (1995) noted that this estimate has a high degree of uncertainty due to
18 inadequate information on the action spectrum and dose-response relationships. Quantitative
19 estimates have not been possible for photokeratitis, pterygium, or other UV-related ocular effects
20 due to lack of epidemiologic and experimental data. As is the case for all of the other UV-
21 related health outcomes, there is no published information on the potential effects of decreased
22 tropospheric O3 concentrations on cataract formation due to any changes in surface-level UV
23 flux resulting from decreases in tropospheric O3.
24
25 10.2.3.5 Ultraviolet Radiation and Immune System Suppression
26 Experimental studies have suggested that exposure to UV radiation may suppress local and
27 systemic immune responses to a variety of antigens (Clydesdale et al., 2001; Garssen and van
28 Loveren, 2001; Selgrade et al., 1997). In rodent models, these effects have been shown to
29 worsen the course and outcome of some infectious diseases and cancers (Granstein and Matsui,
30 2004; Norval et al., 1999). Granstein and Matsui (2004) stated that exposure to UV-B radiation
31 caused immunosuppression in mice ultimately by releasing cytokines that prevent antigen-
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1 presenting cells from performing their normal functions and causing direct damage to epidermal
2 Langerhans cells. Noonan et al. (2003) investigated UV skin cancer induction in two strains of
3 reciprocal Fl hybrid mice and found that genetically determined differences in susceptibility to
4 UV-induced immunosuppression was a risk factor for skin cancer. At high-UV radiation doses,
5 mice with greater susceptibility to immune suppression had a larger proportion of skin tumors
6 compared to those with low susceptibility (Noonan et al., 2003).
7 While results from animal models are supportive of an association between UV radiation
8 and local and systemic immunosuppression, evidence for UV-induced immunosuppression in
9 humans is limited. There is some evidence that UV radiation has indirect involvement in viral
10 oncogenesis through the human papillomavirus (Pfister, 2003). Additional evidence of
11 UV-related immunosuppression comes from an epidemiologic study of 919 patients with rare
12 autoimmune muscle diseases from 15 cities on four continents with variable UV radiation
13 intensity (Okada et al., 2003). Ultraviolet radiation was strongly associated with the prevalence
14 of dermatomyositis, an autoimmune disease distinguished by the presence of photosensitive
15 pathognomonic rashes (Okada et al., 2003). In patients with the human immunodeficiency virus,
16 UV-B radiation lead to activation of the virus in their skin through the release of cytoplasmic
17 nuclear factor kappa B (Breuher-McHam et al., 2001). In a study by Belgrade et al. (2001),
18 UV-induced immunosuppression was examined in 185 subjects with different skin
19 pigmentations. To assess immune suppression, dinitrochlorobenzene was applied to irradiated
20 buttock skin 72 hours after irradiation. Differences in sensitivity were not related to skin type
21 based on the Fitzpatrick classification or minimal erythemal dose (Belgrade et al, 2001).
22 However, erythemal reactivity, assessed by the steepness of the erythemal dose-response curve,
23 was shown to be significantly associated with UV-induced immunosuppression. Only subjects
24 with steep erythemal responses, which included individuals with skin types I through V, showed
25 a dose-response relationship between UV exposure and immune suppression (Belgrade et al,
26 2001).
27 Most action spectrum investigations have concluded that immunosuppression is caused
28 most effectively by the UV-B waveband (Garssen and van Loveren, 2001). The effects of UV-A
29 on local and systemic immunosuppression have been unclear and inconsistent. There is some
30 evidence that high doses of UV-A is protective of immunosuppression induced by UV-B
31 exposure (Halliday et al., 2004). Given the variety of outcomes of immune suppression and
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1 possible mechanisms of effect, little detailed information exists on UV radiation action
2 spectrums and dose-response relationships. The available data are insufficient to develop
3 quantitative risk estimates for UV radiation-induced immunosuppression in humans.
4
5 10.2.3.6 Protective Effects of Ultraviolet Radiation - Production of Vitamin D
6 Most humans depend on sun exposure to satisfy their requirements for vitamin D (Holick,
7 2004). UV-B photons are absorbed by 7-dehydrocholesterol in the skin, leading to its
8 transformation to previtamin D3, which is rapidly converted to vitamin D3. Vitamin D3 is
9 metabolized in the liver, then in the kidney to its biologically active form. Vitamin D deficiency
10 is known to cause metabolic bone disease among children and adults, but also may increase the
11 risk of many common chronic diseases, including type I diabetes mellitus and rheumatoid
12 arthritis (Holick, 2004). Vitamin D also is capable of inhibiting the growth of various cancer
13 cells in cell cultures. Thus, vitamin D deficiency may lead to increased incidence of carcinomas
14 in various organs (Studzinski and Moore, 1995).
15 Large geographical gradients in mortality rates for a number of cancers in the U.S. are not
16 explained by dietary or other risk factors; therefore, it has been hypothesized that some
17 carcinomas are due to insufficient UV-B radiation. For example, published literature indicates
18 that solar UV-B radiation, by increasing production of vitamin D, is associated with reduced risk
19 of cancer of the breast (Freedman et al., 2002; Garland et al., 1990; Gorham et al., 1990; Grant,
20 2002a; John et al., 1999), colon (Freedman et al., 2002), ovary (Freedman et al., 2002; Lefkowitz
21 and Garland, 1994), and prostate (Freedman et al., 2002; Hanchette and Schwartz, 1992), as well
22 as non-Hodgkin lymphoma (Hartge et al., 1996). Eight other cancers, including bladder,
23 esophageal, kidney, lung, pancreatic, rectal, stomach, and corpus uteri, have been found to
24 exhibit an inverse correlation between mortality rates and UV-B radiation (Grant, 2002b).
25 Using UV-B data from July 1992 and U.S. cancer mortality rates from 1970 to 1994,
26 premature cancer deaths attributable to reduced UV-B exposure were analyzed in an ecologic
27 study (Grant, 2002b). The annual number of premature deaths from various cancers due to
28 lower UV-B exposures was 21,700 for white Americans; 1,400 for black Americans; and 500 for
29 Asian Americans and other minorities. Uncertainty in the estimation of UV-B exposure limits
30 the confidence for the estimates of excess cancer deaths attributable to reduced exposure.
31 No study has assessed the decreased risk of cancer mortality resulting from increased UV
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1 radiation attributable to decreased tropospheric O3 levels, but the change in risk is expected to be
2 unappreciable.
3 In establishing guidelines on limits of exposure to UV radiation, the ICNIRP agreed that
4 some low-level exposure to UV radiation benefits health (ICNIRP, 2004). However, the adverse
5 health effects necessitated the development of exposure limits for UV radiation. The ICNIRP
6 recognized the challenge in establishing exposure limits that would achieve a realistic balance
7 between beneficial and adverse health effects.
8
9 10.2.4 Summary and Conclusions for O3 Effects on UV-B Flux
10 Increased solar radiation, especially UV-B radiation, can be detrimental as well as
11 beneficial to the exposed human population. Analogously, both positive and negative effects of
12 UV-B radiation would be expected on plant and animal biota, and on man-made materials (e.g.,
13 Van der Leun et al., 1998; U.S. Environmental Protection Agency, 2002). Other environmental
14 factors include the daily nonlinearity of UV-B flux, nonlinear absorption by O3, and absorption
15 and scattering of UV-B radiation by both stratospheric and tropospheric processes, including
16 ground-level environmental pollutants such as PM and O3. Surface flux variability of UV-B is
17 partially dependent upon cloud and tree cover, latitude, time of year, and time of day. Given the
18 magnitude of this natural variability, it would be extremely difficult with current instrumentation
19 and atmospheric models to measure and/or estimate a change in the ground-level UV-B flux and
20 to attribute it to small reductions in troposphericO3 and/or PM.
21 Of equal or even greater importance are changes in human habits of daily activities,
22 recreation, dress, and skin care. Little is known about the impact of variability in these human
23 factors that can modify individual exposure to UV radiation. Most information about potential
24 UV exposure to the human population relies on demographic information. There is some
25 information available on activity patterns (e.g., EPA NHAPS) and protective behaviors (e.g.,
26 Canadian National Survey on Sun Exposure and Protective Behaviors), but the data are
27 inadequate to accurately assess UV radiation exposure in the population. The uncertainties in
28 characterizing UV radiation exposure limit any UV-related health risk assessment.
29 Another difficulty in examining health risks of UV radiation exposure in the population is
30 due to the lack of detailed information regarding the relevant type (e.g., peak or cumulative) and
31 time period (e.g., childhood or lifetime) of exposure, wavelength dependency of biological
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1 responses, and interindividual variability in UV resistance. For example, exposure to solar UV-
2 B radiation appears to be the most important environmental risk factor for basal and squamous
3 cell carcinomas and cutaneous malignant melanoma in fair-skinned individuals. Originally, total
4 cumulative exposure to UV-B was believed to be the most important environmental factor in
5 determining risk for these tumors. New information now suggests that only squamous cell
6 carcinoma risk is related to total accumulated exposure. For basal cell carcinoma and for
7 malignant melanoma, new information suggests that increases in risk are linked to early
8 exposures (before about age 15), particularly those leading to severe sunburns. There is also
9 controversy regarding the effect of UV-A on health outcomes. Though most studies have found
10 that UV-B radiation is highly mutagenic, some evidence indicates that UV-A radiation, at high
11 doses, may also be carcinogenic.
12 Susceptibility to the various UV radiation-induced health effects, generally, appears to be
13 related to skin type. Numerous studies have shown that individuals with Fitzpatrick skin types I
14 and II (i.e., fair-skinned phenotypes) are at increased risk for erythema and skin cancer.
15 In addition, individuals with a reduced ability to repair DNA have increased susceptibility to
16 both nonmelanoma and melanoma skin cancers. For UV-related immunosuppression, however,
17 it appears that erythemal reactivity, rather than skin type, is a marker for susceptibility.
18 There is some evidence that increased UV radiation exposures resulting from depletion of
19 the stratospheric O3 layer will lead to increases in skin cancer and cataracts. However, the
20 numerous assumptions necessary to calculate these quantitative risk estimates lead to high levels
21 of uncertainty. No study has yet examined the benefits of increased UV radiation exposure due
22 to enhanced production of vitamin D. With the currently available information, the effect of
23 changes in tropospheric O3 on UV-induced health outcomes cannot be critically assessed.
24
25
26 10.3 TROPOSPHERIC OZONE AND CLIMATE CHANGE
27 Water vapor, CO2, O3, N2O, CH4, CFCs, and other polyatomic gases present in the earth's
28 troposphere, trap infrared radiation emitted by the earth's surface, leading to surface warming.
29 This phenomenon is widely known as the "Greenhouse Effect" (Arrhenius, 1896), and the gases
30 involved are known as "greenhouse gases" (GHGs). The term used for the role a particular
31 atmospheric component, or any other component of the greater climate system, plays in altering
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1 the earth's radiative balance is "forcing." In the past decade, the global atmospheric sciences
2 and climate communities have made significant progress in determining the specific role O3
3 plays in forcing climate.
4 The Intergovernmental Panel on Climate Change (IPCC) was founded in 1988 by the
5 World Meteorological Society (WMO) and the United Nations Environmental Program (UNEP)
6 to support the work of the Conference of Parties (COP) to the United Nations Framework
7 Convention on Climate Change (UNFCCC). Drawing from the global climate and atmospheric
8 sciences community for its authors and reviewers, the IPCC produces reports containing
9 thorough assessments of the available, peer-reviewed science regarding the physical climate
10 system, past and present climate, and evidence of human-induced climate change. This section
11 will summarize the reviews of the available information on the forcing properties of tropospheric
12 O3 provided by IPCC Third Assessment Report (IPCC, 200la) will be given, and will describe
13 some of the more recent developments on the subject.
14 The projected effects of global climate change will be briefly explained to provide the
15 context within which O3 serves as a regional, and possibly global, anthropogenic pollutant.
16 The concept of climate forcing is also explained, along with the factors that influence the extent
17 of climate forcing by ozone. The section concludes with a summary of the various estimates that
18 have been placed on the globally-averaged forcing due to ozone.
19
20 10.3.1 The Projected Impacts of Global Climate Change
21 The study of the atmospheric processes involved in global climate change and its potential
22 consequences for human health and global ecosystems is an area of active research. The IPCC
23 Third Assessment Report (TAR) is the most thorough evaluation of current scientific
24 understanding of climate change available at this time. In addition to the first and second IPCC
25 assessments in 1990 and 1995, along with other IPCC reports, earlier assessments included those
26 conducted by the United Nations Environment Program (UNEP, 1986), the World
27 Meteorological Organization (WHO, 1988), the U.S. Environmental Protection Agency (1987),
28 and others (e.g., Patz et al., 2000a,b). The reader is referred to these documents for a complete
29 discussion of climate change science. An abbreviated list of the IPCC conclusions, to date, and a
30 short discussion of the potential impacts of climate change on human health and welfare is
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1 provided here to serve as the context for the discussion of the role of the increasing tropospheric
2 O3 concentration in climate change.
3 According to various historic and modern measurement records, atmospheric GHG
4 concentrations have increased dramatically in the past century, and have been attributed to
5 human activities. The IPCC TAR describes the scientific theory and evidence tying increases
6 in GHGs to human activities (IPCC 2001 a).
7 An increasing body of geophysical observations shows that the earth is warming and that
8 other climate changes are underway. These observations include the global surface temperature
9 record assembled since the year 1860, the satellite temperature record begun in 1979, recorded
10 changes in snow and ice cover since the 1950s, sea level measurements taken throughout the
11 20th century, and sea surface temperature observations recorded since the 1950s. Other
12 evidence includes a marked increase over the past 100 years in the frequency, intensity, and
13 persistence of the zonal atmospheric circulation shifts known as the El Nino-Southern
14 Oscillation (ENSO). ENSO events occur when the tropical Pacific Ocean has accumulated a
15 large, localized mass of warm water that interrupts cold surface currents along South America,
16 altering precipitation and temperature patterns in the tropics, subtropics and the midlatitudes.
17 IPCC (1998, 200la) reports also describe the results of general circulation model (GCM)
18 studies that predict that human activities will alter the climate system in a manner that will likely
19 lead to marked global and regional changes in temperature, precipitation and other climate
20 properties. These changes are expected to increase the global mean sea level, as well as increase
21 the number of extreme weather events including floods and droughts, increased wind speeds and
22 precipitation intensity of tropical cyclones, and changes in soil moisture. These predicted
23 changes can be expected to directly impact human health, ecosystems, and global economic
24 sectors, e.g., hydrology and water resources, food and fiber production (IPCC, 1998, 200Ib).
25 Table 10-1 summarizes these projected impacts.
26 Wide variations in the course and net impacts of climate change in different geographic
27 areas are expected. In general, the projected changes will result in additional stresses on natural
28 ecosystems and human societal systems already impacted by increasing resource demands,
29 unsustainable resource management practices, and pollution. Some of the predicted changes
30 include alterations in ecological balances; in the availability of adequate food, water, clean air;
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Table 10-1. Examples of Impacts Resulting From Projected Changes in Extreme
Climate Events
Projected changes during the 21st Century in
Extreme Climate Phenomena and their
Likelihood3
Simple Extremes
Higher maximum temperatures; more hot days
and heat waves'1 over nearly all land areas (very
likely^)
Higher (increasing) minimum temperatures;
fewer cold days, frost days, and cold waves'1
over nearly all land areas (very likely3)
More intense precipitation events (very likely*
over many years)
Complex Extremes
Increased summer drying over most midlatitude
continental interiors and associated risk of
drought (likely*)
Increase in tropical cyclone peak wind
intensities, mean and peak precipitation
intensities (likely* over some areas)6
Intensified droughts and floods associated with
El Nino events in many different regions
(likely*) (see also under droughts and intense
precipitation events)
Increased Asian summer monsoon precipitation
variability (likely*)
Representative Examples of Projected Impacts*
(all high confidence of occurrence in some areas0)
Increased incidence of death and serious illness in
older age groups and urban poor
Increased heat stress in livestock and wildlife
Shift in tourist destinations
Increased risk of damage to a number of crops
Increased electric cooling demand and reduced energy
supply reliability
Decreased cold-related human morbidity and mortality
Decreased risk of damage to a number of crops, and
increased risk to others
Extended range and activity of some pest and disease
vectors
Reduced heating energy demand
Increased flood, landslide, avalanche, and mudslide
damage
Increased soil erosion
Increased flood runoff could increase recharge of
some floodplain aquifers
Increased pressure on government and private flood
insurance systems and disaster relief
Decreased crop yields
Increased damage to building foundations caused by
ground shrinkage
Decreased water resource quantity and quality
Increased risk of forest fire
Increased risk to human life, risk of infections, disease
epidemics, and many other risks
Increased coastal erosion and damage to coastal
buildings and infrastructure
Increased damage to coastal ecosystems such as coral
reefs and mangroves
Decreased agricultural and rangeland productivity in
drought- and flood-prone regions
Decreased hydro-power potential in drought-prone
regions
Increased flood and drought magnitude and damages
in temperate and tropical Asia
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Table 10-1 (cont'd). Examples of Impacts Resulting From Projected Changes in
Extreme Climate Events
Projected changes during the 21st Century in Representative Examples of Projected Impacts*
Extreme Climate Phenomena and their (all high confidence of occurrence in some areas0)
Likelihood3
Increased intensity of midlatitude storms • Increased risks to human life and health
(little agreement between current models)d • Increased property and infrastructure losses
• Increased damage to coastal ecosystems
"Likelihood refers to judgmental estimates of confidence used by TAR WGI: very likely (90-99% chance); likely
(66-90% chance). Unless otherwise stated, information on climate phenomena is taken from the Summary for
Policymakers, TAR WGI. TAR WGI = Third Assessment Report of Working Group 1 (IPCC, 2001a).
bThese impacts can be lessened by appropriate response measures.
"High confidence refers to probabilities between 67 and 95%.
Information from TAR WGI, Technical Summary.
eChanges in regional distribution of tropical cyclones are possible but have not been established.
Source: IPCC(2001b).
1 and in human health and safety. Poorer nations can be expected to suffer the most, given their
2 limited adaptive capabilities.
3 Although many regions are predicted to experience severe, possibly irreversible, adverse
4 effects due to climate change, beneficial changes may also take place. For example, certain
5 regions may benefit from warmer temperatures or increased CO2 fertilization, e.g., U.S. West
6 Coast coniferous forests, some Western rangelands. Specific benefits may include reduced
7 energy costs for heating, reduced road salting and snow-clearance costs, longer open-water
8 seasons in northern channels and ports, and improved agricultural opportunities in the northern
9 latitudes, as well as the Western interior and coast. For further details about the projected effects
10 of climate change on a U.S.-regional scale, the reader is also referred to several regionally-
11 focused reports (MARAT, 2000; Yarnal et al., 2000; NERAG, 2001; GLRAG, 2000), as well as
12 a report on potential impacts upon human health due to climate change (Bernard et al.,
13 2001).The IPCC report, "The Regional Impacts of Climate Change," (IPCC, 1998) describes the
14 projected effects of human-induced climate change on the different regions of the globe,
15 including Africa, the Arctic and Antarctic, the Middle East and arid Asia, Australasia, Europe,
16 Latin America, North America, the small island nations, temperate Asia, and tropical Asia.
17 While climate models are successful in simulating present annual mean global climate and the
January 2005 10-3 5 DRAFT-DO NOT QUOTE OR CITE
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1 seasonal cycles on continental scales, they have been less successful on regional scales. Clouds
2 and humidity, essential factors in defining local and regional (sub-grid scale) climate, are
3 significantly uncertain (IPCC, 200la). Due to modeling uncertainties, both in reproducing
4 regional climate and in predicting the future economic activity that will govern future GHG
5 emissions, the projected impacts discussed above are also uncertain.
6 Findings from the United States Global Change Research Program (USGCRP) (NAST,
7 2000) report and related reports illustrate the considerable uncertainties and difficulties in
8 projecting likely climate change impacts on regional or local scales. The USGCRP findings also
9 reflect the mixed nature of projected potential climate change impacts (i.e., combinations of
10 mostly deleterious, but other possibly beneficial, effects) for U.S. regions and their variation
11 across different regions. Difficulties in projecting region-specific climate change impacts are
12 complicated by the need to evaluate the potential effects of regional- or local-scale changes in
13 key air pollutants not only on global-scale temperature trends, but also regional- or local-scale
14 temperature and precipitation patterns.
15
16 10.3.2 Solar Energy Transformation and the Components of the Earth's
17 Climate System
18 Mass, in any form, has the capacity to interact with solar and terrestrial radiation, but the
19 manner in which mass interacts with radiation is governed by its particular physical form and/or
20 molecular properties. The most interesting example of radiative properties as a function of
21 physical form is water. In its gas phase, water is the most important GHG present in the climate
22 system, due to its ability to absorb long-wave terrestrial radiation. Conversely, in its frozen form
23 as snow or sea ice, the most important role for water in the climate system is scattering
24 ultraviolet and visible solar radiation back to space, i.e., decreasing the earth's net solar radiation
25 receipts by increasing the earth's reflective properties (albedo). In its liquid aerosol form as
26 clouds, water scatters radiation. In its bulk liquid form as ocean water, it absorbs terrestrial
27 radiation, and represents the earth's most important reservoir of heat energy due to its mass.
28 Molecular properties become important in comparing the radiative absorptive capacity of
29 different gases. In the case of gases, the atomic composition and molecular structure determine
30 the wavelengths of light that a gas molecule can absorb. Ozone and O2, provide one example of
31 the importance of molecular structure in determining absorption capacity. These molecules are
January 2005 10-36 DRAFT-DO NOT QUOTE OR CITE
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1 composed solely of atomic oxygen atoms, but their bond structures are distinct. Ozone, because
2 of its three-atom, bent molecular structure, has the capacity to absorb terrestrial (infrared)
3 wavelengths, making it a GHG. Tropospheric O3, like stratospheric O3, also has the capacity to
4 absorb ultraviolet radiation of 320 nm and shorter, increasing the energy-absorbing capacity of
5 the troposphere. Molecular oxygen, O2, due to its structure, is limited to absorbing very short-
6 wave ultraviolet light - and does so at altitudes too high to influence the climate system
7 significantly.
8 Each component of the climate system plays a role in absorbing, transforming, storing,
9 dispersing, or scattering solar radiation. Weather is a consequence of the transformation and
10 dispersion of terrestrial radiation. The term, "weather," refers to the condition of the earth's
11 atmosphere at a given time and place. It is defined by the air temperature, air pressure, humidity,
12 clouds, precipitation, visibility, and wind speed. The "climate" for a given place on the earth's
13 surface is a long-term average of these elements of weather, accounting for daily and seasonal
14 weather events. The frequency of extreme weather events is used to distinguish among climates
15 that have similar averages (Ahrens, 1994).
16 The earth's capacity for retaining heat is increased by the transformation of its oxygen, for
17 example, into O3 by way of air pollution chemistry. Climate components, including GHGs, land,
18 oceans, sea ice, land ice and snow, atmospheric particles, vegetation, clouds, etc., all contribute
19 to the earth's heat capacity with respect to solar energy. Changes in the properties (or mass) of
20 these components will "force" the climate system in one direction or the other, i.e., warmer
21 versus cooler. Climate forcing is further described, below.
22
23 10.3.3 The Composition of the Atmosphere and the Earth's Radiative
24 Equilibrium
25 The Greenhouse Effect is the term given to the decreased rate of reemission of absorbed
26 solar energy due to the heat-retaining properties of the Earth's atmosphere. According to simple
27 radiative transfer theory, at thermal equilibrium, the earth's temperature should be near -15 °C.
28 This is the temperature of a theoretical "black body" that is receiving and then reemitting
29 342.5 WnT2, i.e., the globally-averaged amount of full-spectrum solar energy absorbed, and then
30 reemitted by the earth as infrared terrestrial radiation, per square meter. In fact, satellite
31 observations well above the atmosphere indicate that the earth's average planetary temperature
January 2005 10-37 DRAFT-DO NOT QUOTE OR CITE
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1 is remarkably close to its theoretical black body value at -18 °C, a temperature at which liquid
2 water ordinarily does not exist.
3 At its surface, however, the earth's average temperature is +15 °C. The +33 °C
4 temperature differential between the Earth's planetary and surface temperatures is due to the
5 presence of infrared radiation-absorbing components in the atmosphere, such as water vapor,
6 CO2, CH4, several other trace gases, and some types of particles and clouds.
7 The atmosphere, when cloud free, is largely transparent in the solar wavelength range.
8 A small fraction of this radiation is absorbed and reemitted as black body radiation by dark
9 atmospheric particles (IPCC, 200la). The majority of clouds and particles offset, in part, the
10 greenhouse effect by increasing the earth's albedo and, therefore, decreasing the overall amount
11 of solar radiation absorbed by the earth system.
12 Ozone, SO2 and NO2 also absorb ultraviolet and near ultraviolet wavelengths, in addition to
13 infrared wavelengths. Once absorbed by a gas molecule, the energy introduced by a photon may
14 induce a photochemical reaction with the residual energy thermally exciting (heating) the
15 products of the reaction. Alternatively, the energy introduced into the molecule by the photon
16 may be dispersed amongst neighboring molecules via intermolecular collisions, or reemitted in
17 part as a lower energy (IR) photon.
18 Radiation from the sun or the earth's surface that is absorbed by gases and particles is
19 reemitted isotropically, i.e., it is equally likely to be emitted in all directions. Therefore, to a
20 first approximation, half of the radiation trapped by the earth's atmosphere is reflected back to
21 its surface. A portion of this radiation is transformed into the heat energy that drives the
22 atmospheric processes that form the basis of weather and climate. Radiation which is not
23 absorbed by gases and aerosols reaches the earth's surface where it is scattered (reflected) or
24 absorbed, depending on the surface albedo.
25 Successful modeling of the earth's climate and, therefore, the assessment of the extent of
26 human-induced climate change and development of appropriate policy depends on the quality of
27 available information on the relative efficiencies, amounts, and spatial and temporal distributions
28 of the various radiatively active components of the atmosphere that absorb and/or reflect solar
29 and terrestrial radiation, along with all the other non-atmospheric components of the earth
30 system.
31
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1 10.3.3.1 Forcing of the Earth's Radiative Balance
2 As mentioned earlier, the commonly used measure of the relative influence of a given
3 component of the climate system on the earth's radiative balance is its radiative forcing (IPCC,
4 2001a; Houghton et al., 1990). Radiative forcing, in WnT2, is a quantity that was developed by
5 the climate modeling community as a first order-only means of estimating relative effects of
6 individual anthropogenic and natural processes on the energy balance within the climate system.
7 Discussions within the climate community are underway regarding a metric to replace forcing as
8 the standard measure of climate impact - one that will account for more of the factors that
9 determine the effectiveness of a given atmospheric component in this capacity. However,
10 forcing remains the current standard (National Research Council, 2005).
11 When the effect of a particular component of the climate system is to reduce the amount of
12 solar energy absorbed, usually by increasing the earth's albedo, this component is said to provide
13 a "negative" forcing, measured in WnT2. The convention assigns a positive value to the forcing
14 induced by climate system components that enhance the greenhouse effect, or otherwise act to
15 increase the heat absorbing capacity of the earth system. Purely reflective atmospheric aerosol,
16 clouds, white rooftops, snow-covered land surfaces, and dense sea ice provide a negative
17 forcing, while highly absorbing dark-colored atmospheric aerosols, GHGs, and increases, due to
18 the melting of sea ice sheets, in dark ocean surfaces positively force the climate system.
19 Global and regional climate are roughly defined by the balance between the large number
20 of positive and negative forcings induced by the many different components of the earth system.
21 However, the earth system responds to these forcings in a complex way, due to feedback
22 mechanisms that are theorized but very difficult to resolve at the quantitative level.
23 A simple example would be the positive feedback associated with melting sea ice. As sea
24 ice melts with increasing surface temperatures, the dark ocean surface is revealed which is more
25 efficient at absorbing infrared radiation, further increasing the rate of warming. The role of
26 feedbacks in determining the sensitivity of climate to changes in radiative forcing is described in
27 detail in the IPCC TAR (IPCC, 200la). In the absence of complete, quantitative, information
28 about climate feedbacks the use of radiative forcing values for the many components of the
29 climate system remains the primary method for estimating their relative importance in climate
30 change.
31
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1
2
3
4
5
6
7
8
9
10
11
12
The IPCC has reported estimated values for forcing by individual radiatively active gases,
and particle-phase components of the atmosphere, derived primarily through expert judgment
incorporating the results of peer-reviewed modeling studies. The forcing estimates, shown in
Figure 10-5, are global averages attributed to known greenhouse gases, including O3; particles;
anthropogenic cirrus clouds; land-use change; and natural solar flux variations. Uncertainty
ranges are assigned to reflect the range of modeled values reported in those studies. The current
estimate of forcing due to long-lived, well-mixed, greenhouse gases accumulated in the
atmosphere from the preindustrial era (ca., 1750) through the year 2000 is + 2.4 WnT2 ± 10%
(IPCC, 2001a). An indication of the level of confidence in each of these estimates is given along
the bottom of this figure, again reflecting the expert judgment of the IPCC.
CD
*j
V
2-
re .g
= E
-------
1 The IPCC reported a global average forcing value of 0.35 ± 0.15 Wrrf2 for tropospheric O3,
2 based on model calculations constrained by climatological observations. The considerations and
3 studies used to estimate this value will be outlined below.
4
5 10.3.4 Factors Affecting the Magnitude of Climate Forcing by Ozone
6 Tropospheric O3 is estimated to have provided the third largest increase in direct forcing
7 since pre-industrial times. It may also play a role in indirect forcing through its participation in
8 the oxidative removal of other radiatively active trace gases, such as CH4 and the HCFCs. Given
9 its relatively short atmospheric lifetime, the distribution of tropospheric O3 is highly variable in
10 geographic extent and time. The direct and indirect forcing it imposes on the climate system,
11 therefore, depends upon its geospatial and temporal distribution, but also depends upon the
12 albedo of the underlying surface. These several variables introduce substantial uncertainty into
13 tropospheric ozone forcing estimates.
14
15 10.3.4.1 Global versus Regional Atmospheric Ozone Concentrations
16 Ozone reacts photochemically at time-scales generally shorter than those for large-scale
17 mixing processes in the atmosphere. Concentrated O3 plumes evolve downwind of strong
18 sources of its precursor pollutants: reactive nitrogen, CO, and non-methane hydrocarbons
19 (NMHCs). The most important of these sources are midlatitude industrialized areas and tropical
20 biomass burning. When viewed from above the atmosphere by satellite-borne spectrometers, O3
21 enhancements appear as relatively localized air masses or regional-scale plumes, usually
22 originating from industrialized areas or areas in which active biomass burning is underway. The
23 IPCC (200la) describes the efforts of several research teams who have analyzed data supplied by
24 the satellite-borne Total Ozone Mapping Spectrometer (TOMS) and other remote-sensing
25 instruments to map the global distribution of tropospheric O3 and to attempt to identify processes
26 that influence the global tropospheric O3 budget (IPCC, 200la). More recently, coincident
27 observations of total O3 by TOMS and the Solar Backscattered UV (SBUV) instrument were
28 used by Fishman et al. (2003) to construct well-resolved spatial and temporal maps of the
29 regional distribution of tropospheric O3. Their results were consistent with those reported by
30 others, but with higher regional-scale resolution. They reported large O3 enhancements in the
31 southern tropics in austral Spring, and in the northern temperate latitudes in the summer. The
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1 regional nature of high O3 concentrations was clearly visible in northeastern India, the eastern
2 United States, eastern China, and west and southern Africa, each coincident with high population
3 densities. Fishman et al. (2003) noted, as have the other groups cited above, significant
4 interannual variability in the concentrations observed over these regions. In situ measurements
5 of tropospheric O3 concentrations range from 10 ppb over remote oceans, to 100 ppb in both the
6 upper troposphere and in plumes downwind from polluted metropolitan regions (IPPC, 200la).
7 Ground-level concentrations in urban areas are often higher than 100 ppb. In the southern
8 hemisphere, one of the largest sources of O3 precursors is biomass burning. Biomass burning
9 elevates O3 on large spatial scales, particularly in the tropical Atlantic west of the coast of Africa
10 and in Indonesia.
11 Current estimates place the global burden of tropospheric O3 at a highly uncertain
12 370 Tg, equivalent to an average column density of 34 Dobson Units (1 DU = 2.687 x 1016
13 molecules/cm"2) or a mean concentration of 50 ppb (IPCC, 200la). Accounting for differences
14 in levels of industrialization between the hemispheres, the average column burden in the
15 Northern Hemisphere is estimated to be 36 DU, with the Southern Hemisphere estimated to
16 average 32 DU. Due to its rapid photochemistry, individual surface measurements of
17 tropospheric O3 cannot capture large scale concentrations, nor will it represent the concentration
18 at higher altitudes. Dense surface and vertical measurements (ozonesondes) would be required
19 to supplement available output from remote sensing instruments to provide the complete set of
20 observations needed to derive a credible global O3 budget. Such a measurement program
21 appears, at present, to be impractical.
22 Little historical data exists that might be used to estimate the global ozone burden prior to
23 industrialization. Although a few late 19th century measurements suggest that O3 has more than
24 doubled in Europe during the 20th century. The insufficient data record on pre-industrial
25 tropospheric O3 distributions introduces a major uncertainty in the estimation of the ozone-
26 induced forcing (IPCC, 200la).
27
28 10.3.4.2 Global Versus Regional Atmospheric Ozone Trends
29 For the northern hemisphere, weekly continuous data since 1970 are available from only
30 nine stations in the latitude range 36°N to 59°N (IPCC, 2001a). Available tropospheric O3
31 measurements do not reveal a clear trend in concentration, while trends in the stratosphere are
January 2005 10-42 DRAFT-DO NOT QUOTE OR CITE
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1
2
3
4
5
6
7
8
9
10
more readily identified. Different trends are seen at different locations for different periods,
consistent with regional changes in pollutant emission, especially NOX. Logan et al. (1999)
analyzed the composite record of mid-tropospheric ozone abundance from the nine station
network. A plot of data is shown in Figure 10-6. While no clear trend appeared for 1980
through 1996, the second half of this record (about 57 ppb) is clearly greater than the first half
(about 53 ppb). The trend may be consistent with changes in regional NOx emission rates due to
pollution reduction efforts in developed countries and increasing emissions in rapidly growing
economies in Asia.
70
65
60
2"
Q.
S 55
O
50
45
40
400 to 630 hPa
36°N to 59°N
1970
1975
1980
1985 1990 1995
Figure 10-6. Mid-tropospheric O3 abundance (ppb) in northern midlatitudes
(36 °N-59 °N) for the years 1970 to 1996. Observations between 630 and
400 hPa are averaged from nine ozone sonde stations (four in North
America, three in Europe, two in Japan), following the data analysis of
Logan et al. (1999). Values are derived from the residuals of the trend fit
with the trend added back to allow for discontinuities in the instruments.
Monthly data (points) are shown with a smoothed 12-month-running
mean (line).
Source: IPCC(2001).
1 It must be noted that the measurements shown in Figure 10-6 are for surface
2 concentrations, only. Many fewer locations have measured changes in the concentrations of O3
3 as a function of altitude. Fewer still are locations that have collected and maintained data
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1 records prior to 1970. The absence of historical data on the vertical distribution of O3 adds to the
2 difficulty in estimating historical atmospheric burdens and trends in O3-related climate forcing.
3 The IPCC (200la) surveyed the results of published chemistry transport model (CTM)
4 modelling studies, listed in Table 10-2, that estimated the global average increase since the
5 pre-industrial era in total column O3. Model estimates ranged from +7 to +12 DU. On the basis
6 of these estimates, available measurements and other analyses, the IPCC estimated that total
7 column O3 has increased by 9 DU, with a 67% confidence range of+6 to +13 DU. In some of
8 the modelling studies, emissions scenarios were used that predict a further increase in column
9 O3 due to growing emissions of O3 precursors.
10
11
Table 10-2. CTM Studies Assessed by the IPCC for its Estimate of the Change in
Global, Total Column O3 Since the Pre-industrial Era
Estimated Change in Column
Ozone in DU
7.9
8.9
8.4
9.5
12
7.2
8.7
9.6
8
Model Used
GFDL
MOZART-1
NCAR/2D
GFDL-scaled
Harvard/GISS
ECHAM4,
UKMO,
UIO
MOGUNTIA
Authors (Publication Date)
Haywoodetal. (1998)
Hauglustaine et al. (1998)
Kiehletal. (1999)
Levyetal. (1997)
Mickley et al. (1999)
Roelofsetal. (1997)
Stevenson et al. (2000)
Berntsen et al. (1999)
VanDorland et al. (1997)
Source: IPCC (2001).
1 Chapters 2 and 3 of this document provide an expanded discussion of the issues associated
2 with determining the global tropospheric O3 background.
3
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1 10.3.4.3 The Sensitivity of Ozone-related Forcing Surface to Albedo
2 The characteristics of the surface underlying an O3 enhancement play a large role in the O3
3 forcing effect. Highly reflective, warm surfaces, such as light-colored deserts, scatter solar
4 ultraviolet (UV) radiation; absorb, and then emit infrared terrestrial radiation. Both forms of
5 radiation can be trapped, transformed, and/or reemitted back to the surface by tropospheric O3.
6 Highly reflective, cold, surfaces will scatter more radiation while emitting less terrestrial infrared
7 radiation. Dark, warm, surfaces such as tropical ocean hot spots predominantly emit terrestrial
8 radiation. Ozone will trap heat at differing efficiencies as a consequence of the amount and type
9 of radiation reflected or emitted from the surface underneath it. Studies by two groups,
10 Hauglustaine et al. (1998) and Mickley et al. (1999), have shown that industrial pollution that
11 has been transported to the Arctic induces a high, regional O3-related forcing due to the highly
12 reflective underlying ice and snow surface.
13
14 10.3.5 Estimated Forcing by Tropospheric Ozone
15 10.3.5.1 Direct Climate Forcing Due to Ozone
16 The inhomogeneous distribution of O3 within the troposphere coupled with the large
17 uncertainty in the global O3 budget significantly complicates the matter of estimating the global
18 average direct forcing due to O3. The IPCC TAR (200la) lists the results of several modeling
19 studies that estimated the annual change in the relative forcing by O3 since pre-industrial times.
20 It was noted that the differences amongst the estimates were most likely due to differences in
21 predicted O3 chemistry, including the emissions inventories used, and the chemical process and
22 transport mechanisms incorporated into the models, rather than by factors relating to radiative
23 transfer. The IPCC intercomparison of the models and their results indicated that the
24 uncertainties in estimated forcings due to O3 have reduced since the IPPC Second Assessment
25 Report (1996)
26 The O3-related forcings estimated by studies considered by the IPCC (200la) are listed in
27 Table 10-3. Ten of the listed estimates are based on global chemistry/transport models
28 calculations. One study was constrained by a climatology derived from observations. Given the
29 differences in calculated total column O3 amongst the models, a normalized forcing (WnT2 per
30 Dobson Unit of tropospheric O3 change) is listed in addition to the absolute forcing (WnT2)
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Table 10-3. Tropospheric O3 Change (AO3) in Dobson Units (DU) Since Pre-industrial
Times, and the Accompanying net (SW plus LW) Radiative Forcings (Wm~2),
After Accounting for Stratospheric Temperature Adjustment (using the Fixed
Dynamical Heating method). Estimates are Taken From the Published Literature.
Normalized Forcings (norm.) Refer to Radiative Forcing per O3 Change (Wm~2 per DU)
Estimated Global Average Forcing Due to TronosDheric Ozone
Clear sky conditions
Reference
Berntsen et al. (1997) - [Reading model]
Stevenson etal. (1998)
Berntsen et al. (1997) - [Oslo model]
Haywoodetal. (1998a)
Kiehl etal. (1999)
Berntsen et al. (2000)
Brasseur etal. (1998)
van Dorland et al. (1997)
Roelofs etal. (1997)
Lelieveld and Dentener (2000)
Hauglustaine et al. (1998)
Mean
iO3
7.600
8.700
7.600
7.900
8.400
9.600
8.070
7.200
—
8.940
8.224
Net
0.310
0.391
0.390
0.380
0.379
0.428
0.443
0.397
—
0.511
0.403
Net (norm.)
0.041
0.045
0.051
0.048
0.045
0.045
0.055
0.055
—
0.057
0.049
Total sky conditions
Net
0.280
0.289
0.310
0.310
0.320
0.342
0.370
0.380
0.404
0.420
0.426
0.343
Net (norm.)
0.037
0.033
0.041
0.039
0.038
0.036
0.047
0.056
—
0.048
0.042
Source: IPCC(2001).
1 estimated by each model. Both clear sky (cloud-free) and total sky (including clouds) forcing
2 estimates are listed.
3 The largest O3-related forcings coincide with strongest sources of tropospheric ozone, the
4 models predict that occur in the northern midlatitude regions (40° to 50° N), reaching as much as
5 1 WnT2 in the summer and in the tropics, related to biomass burning. In general, the estimates
6 are comparable in magnitude and show similarity in geographic distribution. For total sky
7 conditions, the range in globally and annual averaged tropospheric O3 forcing from all of these
8 models is from 0.28 to 0.43 WnT2, while the normalized forcing is 0.033 to 0.056 WnT2 per DU.
9 As expected, they are opposite in sign to the forcing estimated for sulphate aerosols, which
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1 scatter radiation. The range in normalized forcings emphasizes the differences in assumptions
2 used by the different models. The tropospheric O3 forcing constrained by the observational
3 climatology is 0.32 WnT2 for globally averaged, total sky conditions. As shown in Figure 10-1,
4 the IPCC (2001) concluded that 0.35 ± 0.15 WnT2represents the most likely value for annually
5 and globally-averaged forcing by tropospheric O3.
6
7 10.3.5.2 Indirect Forcing Due to Ozone
8 Ozone has an indirect climate forcing effect due to its role in the oxidative removal of
9 other reactive GHGs, including CH4, hydrofluorocarbons (HFCs) and other reactive
10 non-methane hydrocarbons (NMHCs). The primary actor in this effect is a second generation
11 product of the photolysis of O3, the hydroxyl radical (OH). Hydroxyl radical is produced by
12 way of a pair of reactions that start with the photodissociation of O3 by solar UV.
13
14
O3 + hv ->• O(1D) + O2 (10-5)
16
17 O(1D) + H2O -> OH + OH (10-6)
18
19 Reactions with OH are the primary removal mechanism for CH4 and NMHCs as well as the
20 pollutants NOX and CO. Methane and CO have especially high abundances in the global
21 atmosphere. OH is estimated to react with these two gases within one second of its formation.
22 In addition to CH4, NOX, CO and the NMHCs, OH concentrations are controlled by local
23 concentrations of H2O (humidity) and the intensity of solar UV. Different atmospheric
24 concentrations of the required precursors suggest that pre-industrial OH concentrations are likely
25 to have been different from present-day concentrations, but there is no consensus on the
26 magnitude of this difference. Observations of global atmospheric concentrations of chloroform
27 (CH3CC13), a well-mixed tropospheric species that reacts with OH, have been used to estimate
28 OH abundances. Independent studies have shown overlapping trends for the period 1978 to
29 1994, but none outside the given uncertainty ranges (0.5 ± 0.6%/yr) (Prinn et al., 1995; Krol
30 et al., 1998). The IPCC (2001) reported a range of+5% to -20% for predicted changes in global
31 OH abundances.
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1 Given the difficulty in estimating global OH abundances in the past, present and future,
2 estimates of indirect forcing due to O3 have been difficult to obtain and are highly uncertain.
3 Attempts have been made to account for OH feedbacks in estimating the global warming
4 potential of CH4.
5
6 10.3.5.3 Predictions for Future Climate Forcing by Anthropogenic Ozone
7 The rate of increase in surface O3 in Europe and North America, since 1980, appears to be
8 slowing, likely due to control measures intended to improve urban air quality. Not surprisingly,
9 CTM modeling attempts to predict future precursor emissions and resulting O3 abundances
10 indicate that the largest future O3-related forcings will be related to population growth and
11 economic development in Asia (van Borland et al., 1997; Brasseur et al., 1998). The results of
12 these modeling studies suggested that a higher globally averaged total radiative forcing due to O3
13 from pre-industrial times to 2050 of 0.66 WnT2 and 0.63 WnT2. Chalita et al. (1996) predicted a
14 globally averaged radiative forcing from pre-industrial times to 2050 of 0.43 WnT2. Stevenson
15 et al. (1998) predicted an O3-related forcing of 0.48 WnT2 in 2100. All of these predictions must
16 be viewed with much caution given the considerable uncertainties associated with such
17 estimates.
18
19 10.3.6 Conclusion
20 The general consensus within the atmospheric sciences community, as represented by the
21 United Nations Intergovernmental Panel on Climate Change (IPCC), is that human activities
22 have a discernable effect on the earth's climate. However, quantifying the extent of human-
23 induced forcing on climate requires detailed information about human-induced change on the
24 components of the earth system that govern climate. Troposphere ozone is a well-known GHG,
25 but information regarding its historical trends in concentration, its current and future
26 atmospheric burden, and other critical details needed for estimating its direct and indirect
27 forcing effects on the climate system are highly uncertain.
28 The IPCC has estimated that the globally averaged forcing due to O3 is approximately
29 0.35 ± 0.15 WnT2. The role of O3 in climate is likely to be much more pronounced adjacent to
30 the sources of its chemical precursors, consistent with satellite observations of high regional
31 scale column densities near large urban areas and large-scale biomass burning activity.
January 2005 10-48 DRAFT-DO NOT QUOTE OR CITE
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1 Modeling studies evaluated by the IPCC have estimated that regional scale forcing due to O3 can
2 approach 1 WnT2, or two-thirds the forcing estimated for CO2. However, more reliable
3 estimates of the overall importance of forcing due to tropospheric O3 await further advances in
4 monitoring and chemical transport modeling.
5
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14
15
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i 11. EFFECT OF OZONE ON MAN-MADE MATERIALS
2 Ozone and other photochemical oxidants react with many of the economically important
3 man-made materials, decreasing their useful life and aesthetic appearance. Some of the
4 materials known to be damaged by ozone include elastomers, fibers and dyes, and paints. This
5 section will provide a brief discussion on the effects of ozone on man-made materials including
6 the damage mechanisms and, where possible, concentration-response relationships. Since only
7 limited information has been published on the effects of ozone on materials, this section will
8 provide a summary of information presented in the previous ozone criteria document (U.S.
9 Environmental Protection Agency, 1996) and a more detailed discussion of studies published
10 since publication of the previous ozone criteria document. The reader is referred to the previous
11 ozone criteria document for a more detailed discussion of the earlier studies.
12
13 11.11.1 Mechanisms of Ozone Damage and Exposure-Response
14 Elastomer Cracking
15 The elastomeric compounds, natural rubber and synthetic polymers and copolymers of
16 butadiene, isoprene, and styrene, are particularly susceptible to even low levels of ozone.
17 Elastomeric compounds are long chain unsaturated organic molecules. Ozone damages these
18 compounds by breaking the molecular chain at the carbon-carbon double bond; a chain of three
19 oxygen atoms is added directly across the double bond, forming a five-membered ring structure
20 (Mueller and Stickney, 1970). The change in structure promotes the characteristic cracking of
21 stressed/stretched rubber called "weathering." A 5% tensile strain will produce cracks on the
22 surface of the rubber that increase in number with increased stress/stretching. The rate of crack
23 growth is dependent on the degree of stress, the type of rubber compound, ozone concentration,
24 time of exposure, ozone velocity, and temperature (Bradley and Haagen-Smit, 1951; Lake and
25 Mente, 1992) (Gent and McGrath, 1965). Once cracked, there is further ozone penetration and
26 in additional cracking and eventually mechanical weakening or stress relaxation (U.S.
27 Environmental Protection Agency, 1996). Razumovskii et al. (1988) demonstrated the effect of
28 ozone on stress relaxation of polyisoprene vulcanizates. A decrease in stress (stress relaxation)
29 was caused by ozone-induced cracks in exposed elastomers resulting in irreversible changes in
30 the elastomer dimensions and decreased tensile strength. A description of findings of earlier
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1 studies appear in the 1996 Air Quality Criteria Document for ozone (U.S. Environmental
2 Protection Agency, 1996).
3 To counteract the effect of ozone on elastomers, antiozonants and wax have been added to
4 the elastomeric formulations during processing. An antiozonant is an additive used to protect a
5 polymer against the effects of ozone-induced degradation and, hence, used mainly in diene
6 rubbers. Antiozonant protection works either by providing a physical barrier to ozone
7 penetration forming a thin surface film of an ozone-resisting wax or by chemically reacting with
8 ozone or polymer ozonolysis products, as do aromatic diamines such as p-phenylene diamine
9 derivatives. The antiozonant diffuses to the surface of the elastomeric material where it reacts
10 with ozone faster than ozone reacts to break the molecular chain and the carbon-carbon double
11 bond, or the antiozonant diffuses to the surface of the material but is not reactive with ozone and
12 serves as a protective coating against ozone attack. The antiozonant may also serve to scavenge
13 ozone while also providing protective film against ozone attack (Andries et al., 1979; Lattimer
14 etal., 1984).
15 Most of the studies on ozone effects on elastomers were designed to evaluate the
16 effectiveness of antiozonants in counteracting the rubber cracking produced by ozone exposure.
17 Consequently, many of the studies were conducted using ozone concentrations higher than those
18 typically found in the ambient air. Natural rubber strips exposed to high concentrations of ozone
19 (20,000 ppm) under stressed conditions cracked almost instantaneously and were broken within
20 1 sec. When the ozone concentration was lowered (0.02 to 0.46 ppm), the time to required to
21 produce cracks in the exposed rubber material was increased (Bradley and Haagen-Smit, 1951).
22 Lake and Mente (1992) studied the effect of temperature on ozone-induced elastomer cracking
23 and antiozonant protection on natural rubber, epoxidised natural rubber, and two acrylonitrile-
24 butadiene copolymers under constant strain. Temperatures ranged from -20° C to +70° C. The
25 elastomers were exposed to 0.05 to 1,000 ppm ozone for 16 h. Ozone cracking decreased at
26 lower ambient temperatures, however, diffusing of both chemical and wax antiozonants also
27 were slowed at the lower temperatures. Cracking was slightly increased at the higher
28 temperatures but the antiozonants offered more protection.
29 Serrano et al. (1993) evaluated the appropriateness of using ozone-induced elastomer
30 cracking to estimated the ambient ozone concentrations. Two vulcanized natural rubber
31 compounds were exposed for 24 h to varying ozone concentrations under stressed conditions.
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1
2
3
4
5
6
7
Ozone concentrations were 60, 80, 90, 100, and 120 ppb for durations of 2, 4, or 6 h. The 24 h
average ozone concentrations ranged from 31 to 57.5 ppb. There was a clear relationship
between the 24-h average ozone concentration and the distribution of crack length frequencies
on the rubber surface. Table 11-1 gives the average 24-h ozone concentration and lengths for
two vulcanized natural rubber strips.
Table 11-1. Average 24-h Ozone Concentrations Producing the Highest Frequency of
Cracks of a Certain Length in the Middle and Central Zones of the Rubber Test Strips
1% Antiozonant 4010NA #
Crack Length (mm)
0.05-0.10
0.10-0.15
0.15-0.20
0.20 - 0.40
Middle Zones
37.5
45.0
48.0
> 57.5
Central Zones
37.5
48.0
> 57.5
> 57.5
0.5% Antiozonant 4010NA
Middle Zones
40.0
48.0
> 57.5
> 57.5
Central Zones
42.5
53.0
> 57.5
> 57.5
Ozone concentrations given in ppb.
Adapted from Serrano et al. (1993).
1 11.11.2 Textiles and Fabrics
2 Ozone can damage textiles and fabrics by methods similar to those associated with
3 elastomers. Generally, synthetic fibers are less affected by ozone than natural fibers, however,
4 ozone contribution to the degradation of textiles and fabrics is not considered significant (U.S.
5 Environmental Protection Agency, 1996). A study reported by Bogaty et al. (1952) showed that
6 ozone effects moistened cloth more than dry cloth. Scoured cotton duck cloth and commercially
7 bleached cotton print cloth were exposed to 20 to 60 ppb for 1,200 h (50 days). The rate of
8 deterioration was measured by the changes in cuprammonium fluidity values and the fabric
9 breaking strength. At the end of the 1,200-h exposure, there was a 20% loss in breaking
10 strength. Table 11-2 list the changes in cuprammonium fluidity values for both fabrics.
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Table 11-2. Cuprammonium Fluidity of Moist Cotton Cloth Exposed
to 20 to 60 ppb Ozone
Duration of Exposure (h) Cuprammonium Fluidity (rhes)
Duck Cloth 0 2.6
200 2.8
680 4.0
960 6.8
1200 9.5
Bleached Print Cloth 0 8.2
200 8.7
510 9.4
650 12.0
865 12.7
1500 16.5
Adapted from Bogaty et al. (1952).
1 11.11.3 Dyes, Pigments, and Inks
2 Ozone fading of textile dyes is diffusion-controlled; the rate of fading is controlled by the
3 diffusion of the dye to the fiber surface. Many textile dyes react with ozone; however, the rate
4 and severity of the ozone attack is influenced by the chemical nature of the textile fiber and the
5 manner in which the dye is applied. Ozone molecules break the aromatic ring portion of the dye
6 molecule, oxidizing the dye (U.S. Environmental Protection Agency, 1996). In case of aromatic
7 azo dyes, ozone attacks the aromatic rings and electron rich nitrogen atoms (Matsui et al., 1988).
8 Grosjean et al. (1987; 1988a,b) proposed a mechanism of reaction of ozone and indigo,
9 thioindigo, and dibromoindigo, alazarin, and curcumin dyes under dark conditions. Ozone
10 attaches to the dye molecule at the unsaturated carbon = carbon bond. An ozone adduct is
11 formed (1,2,3-trioxolane), followed by scission of the carbon-carbon bond and the subsequent
12 formation of the corresponding Criegee biradical. A similar mechanism was proposed for the
13 reaction of ozone with triphenylmethane colorant Basic Violet 14. Ozone attacked Basic Violet
14 14 at the carbon=carbon unsaturated bond and at the carbon-nitrogen unsaturated bond under
15 dark conditions. Other members of the group of triphenylmethane colorants with unsaturated
16 carbon-carbon bonds also are expected to be subject to ozone fading. Tripheylmethane
17 colorants that are expected to be ozone-fugitive include the amino-substituted cationic dyes
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1 (Malachite Green, Brilliant Green, Crystal Violet, Pararosaniline Chloride, Methyl Green, and
2 others) (Grosjean et al., 1989).
3 An indication that ozone caused textile dye fading was first reported by Salvin and Walker
4 (1955). The researchers found that the fading was primarily the result of the destruction of the
5 blue dye molecule. Drapes made of acetate, Arnel, and Dacron and dyed with anthraquinone
6 blue dye exhibited a decrease in shade that was not accompanied by the characteristic reddening
7 caused by NOX. Figures 11-1 and 11-2 demonstrate the effect of ozone exposure on nylon 6 yarn
8 colored with several blue dyes. Nylon samples inside the home were located on a wall away
9 from sunlight. Outside nylon samples were placed on a covered patio or under the eaves of the
10 house to minimize exposure to sunlight and rain. Ozone concentrations ranged from 2 to 5 ppb
11 outside and 0 to 2 ppb inside. The percent change in dye color was determined monthly by
12 extraction and analysis of the remaining dye or by instrumental measurement of the color change
13 (Haylock and Rush, 1978).
14
15 11.11.4 Artists' Pigments
16 Several artists' pigments are sensitive to fading and oxidation by ozone when exposed to
17 concentrations found in urban areas (Shaver et al., 1983; Drisko et al., 1985; Whitmore et al.,
18 1987; Whitmore and Cass, 1988; Grosjean et al., 1993). The organic pigments that are ozone
19 fugitive include alizarin red pigments containing lakes of the poly cyclic aromatic compound
20 1,2-dihydroxyanthraquinone, blue-violet pigments containing substituted triphenylmethane
21 lakes, indigo, and yellow coloring agents containing polyfunctional, polyunsaturated compounds
22 such as curcumin (Grosjean et al., 1987). Because of the potential of ozone to damage works of
23 art, recommended limits on ozone concentrations in museums, libraries, and archives are
24 relatively low, ranging from 0.013 to 0.01 ppm.
25 Experimental studies demonstrate a concentration x time (C x T) relationship between
26 ozone concentration and exposure time and pigment fading. Cass et al. (1991) summarized
27 some of the earlier research on the effects of ozone on artists' pigments. In studies evaluating
28 the effect of ozone on organic and inorganic watercolors and traditional organic pigments only
29 the traditional organic pigments showed measurable fading from ozone exposure. Of the
30 inorganic pigments tested, only the arsenic sulfides showed ozone-related changes. The
pigments were exposed to 0.3 to 0.4 ppm ozone for 3 mo in the absence of light, at 22 °C and
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100 r
J FMAMJ J ASOND J
Months
Figure 11-1. In-service fading of nylon 6 yarn inside house. • = C.I. Disperse Blue 3;
O = C.I. Basic Blue 22; A = C.I. Acid Blue 27; x = C.I. Disperse Blue 56;
A = C.I. Acid Blue 232.
Source: Haylock and Rush (1978).
1 50% RH. The authors equated this exposure to a C x T of 6 to 8 years inside a Los Angeles
2 museum with air conditioning but without a pollutant removal system.
3 Whitmore and Cass (1988) studied the effect of ozone on traditional Japanese colorants.
4 Most of these compounds are insoluble metal salts that are stable in light and air. Suspensions or
5 solutions of the colorants were airbrushed on hot-pressed watercolor paper or silk cloths.
6 A sample of Japanese woodblock print also was included in the analysis. Samples were exposed
7 to 0.4 ppm ozone at 22 °C, 50% relative humidity, in the absence of light for 12 wk. Changes in
8 reflectance spectra were used to evaluate the effect of ozone exposure on colorant fading.
9 Among the colorants tested on paper, curmin, indigo, madder lake, and lac lake were the most
10 sensitive to ozone exposure. Gamboge was relatively insensitive to ozone.
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100 r
90 -
J FMAMJ J ASOND J
Months
Figure 11-2. In-service fading of nylon 6 yarn outside house. • = C.I. Disperse Blue 3;
O = C.I. Basic Blue 22; A = C.I. Acid Blue 27; x = C.I. Disperse Blue 56;
A = C.I. Acid Blue 232.
Source: Haylock and Rush (1978).
1 The blue and green areas of the sample from the woodblock print was very reactive due to the
2 indigo dye ozone sensitivity. The other colorants, red, yellow, and purple, showed very little
3 sensitivity to ozone. The textiles dyes that reacted with ozone were indigo, alone or in
4 combination with several yellow dyes.
5 Ye et al. (2000) reported the rate of ozone fading of traditional Chinese plant dyes. Twelve
6 different colorants were applied to watercolor paper and silk and exposed to 0.4 ppm ozone at
7 25° C, at 50% RH, in the absence of light for 22 wks. Dye fading was greater when the colorant
8 was applied to the watercolor paper compared to the silk cloth due to the darker initial depth of
9 the shade, the greater saturation of the colorant throughout the cloth. Turn eric, gromwell, and
10 violet on paper was particularly reactive. Tangerine peel was moderately reactive and sappan
11 wood, dalbergia wood, Chinese gall, indigo, and Chinese yellow cork tree were slightly reactive
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1 to ozone. Black tea was not reactive to ozone. The colorants on silk samples showing color
2 changes were gromwell, sappan wood, gardenia, tummeric, and violet. Figures 11-3 and 11-4
3 demonstrate the color change of the various colorants on watercolor paper and silk.
4 Artists' pigments also have exhibited fading when exposed to a mixture of photochemical
5 oxidants. Grosjean et al. (1993) exposed 35 artists' pigments to a mixture of photochemical
6 oxidants consisting of ozone, nitrogen dioxide (NO2), and peroxyacetyl nitrate (PAN) for
7 12 wks. Weekly average photochemical concentrations were 200 ppb for ozone, 56 ± 12 to
8 99 ± 24 for NO2, and Il±3tol8±2 for PAN. All exposures were carried out at room
9 temperature in the absence of light. To determine the effect of humidity on pigment fading, the
10 relative humidity was increased from 46% after 8 weeks of exposure to 83% for a 2 week period
11 and then returned to 46% for the remainder of the exposure.
12 Table 11-3 lists the artists' pigment and degree of fading. Eleven of the pigments
13 exhibited negligible color change, 12 had small color changes, 3 had modest color changes, and
14 9 exhibited substantial color changes. Fading of Disperse Blue 3 and Reactive Blue 2 were
15 likely the result of NO2 exposure, the fading of triphenylmethanes is consistent with exposure to
16 nitric acid formed under high humidity conditions. Fading of the indigos was dominated by
17 ozone exposure and curcumin was faded by all of the photochemicals studied. Increasing the
18 relative humidity resulted in a substantial color change for all of the pigments, with the
19 exception of curcumin and indigo.
20
21 11.11.5 Surface Coatings
22 Ozone will act to erode some surface coatings (paints, varnishes, and lacquers). However,
23 many of the available studies on ozone degradation of surface coatings do not separate the
24 effects of ozone from other pollutants or environmental factors such as weather, humidity, and
25 temperature. Campbell et al. (1974) attempted to demonstrate an ozone related effect on oil
26 house paint, acrylic latex coating, alkyd industrial maintenance coating, urea alkyd coil coating,
27 and nitrocellulose/acrylic automotive paint. Painted test panels were exposed to 100 and
28 1,000 ppb ozone in a xenon arc accelerated weathering chamber for up to 1,000 h. Using weight
29 loss as a measure of ozone-induced erosion the researchers concluded that all of the paints tested
30 suffered degradation in the presence of ozone and that the automotive finish suffered the most
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35
30
25
0)
o 20
15
o
o
o
10
zi cao (gromwe
zi ding cao (violet)
5 10 15 20
Weeks of Exposure to 0.40 ppm Ozone
very reactive colorants
e 3
o
o
o
-0- su mu (sappan wood)
-*- Jiang xiang (dalbergia wood)
-Ar hong chaye (black tea)
. -Oh wu bei zi (Chinese gall)
-O- ban Ian gen (indigo)
"V~ ju zi pi (tangerine peel)
"A- huang bai (Chinese yellow cork tree)
0 5 10 15 20
Weeks of Exposure to 0.40 ppm Ozone
moderately reactive colorants
Figure 11-3. Observed color changes for natural colorant-on-paper systems during
exposure to 0.40 ppm ozone at 25 °C ± 1 °C, 50% RH, in the absence of light.
Source: Ye et al. (2000).
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a)
1
D
o 2
o
O
. su mu (sappan wood)
- zi cao (gromwell)
zi ding cao (violet)
- Jiang huang (tumeric)
- huang zhi zi (gardenia)
J =
10
15
20
Weeks of Exposure to 0.40 ppm Ozone
reactive colorants
2.0
1.5
HI
-------
Table 11-3. Color Change After 12 Weeks of Exposure to a
Mixture of Photochemical Oxidants
Colorant*
Color Change (A£" units)f
Chemical Functionality or
Chemical Composition
Acid Red 37 (17045)*
Acid Yellow 65 *
Alizarin Carmine
Alizarin Crimson
(Pigment Red 83)
Aurora Yellow (77199)
Basic Fuschin (42510)*
Brilliant Green (42040)*
Brown Madder
Cadmium Yellow (77199)
Carmine
Chrome Yellow (77600)*
Copper phthalocyamne
(Pigment Blue 15)
Crimson Lake
Curcumin (Natural Yellow 3)
Disperse Blue 3
French Ultramarine Blue
Gamboge (Natural Yellow 24)
Hooker's Green Light
Indigo (a formulation)
Indigo carmine *
Indigo (73000)*
Mauve
New Gamboge
11.7 ±0.5
1.8 ±0.5
1.8 ±0.2
1.4 ±0.2
0.5 ±0.1
33.4 ±3.0
20.6 ±2.1
1.7±0.1
0.4 ±0.1
1.8 ±0.2
1.7 ± 1.2
l.OiO.l
3.5 ±0.3
15.2 ±2.6
10.8 ±0.1
0.8 ±0.3
0.4 ±0.1
1.5 ±0.4
l.liO.l
14.0 ±1.9
64.1 ±4.5
3.6 ±0.5
0.9 ±0.1
Aminophenyl-substituted azo dye,
sulfonate salt
Nitro- and phenyl-substituted azo dye,
sulfonate salt
Alizarin lake
Alizarin lake
Cadmium sulfide
Amino-substituted triphenylmethane
Amino-substituted triphenylmethane
Alizarin lake
Cadmium sulfide
Lake of cochineal (substituted
anthraquinone)
Lead chromate
Copper phthalocyanine
Alizarin lake
1,7 bis (4-hydroxy-3-methoxyphenyl)-
1,6-heptadiene-3,5-dione
Amino-substituted anthraquinone
Gambogic acid
Chlorinated copper phthalocyanine plus
ferrous beta naphthol derivative
Alizarin lake plus lampblack plus copper
plthalocyanine
5,5-indigo disulfonic acid, sodium salt
Lake of triphenyl methane (basic fuschin)
plus copper phthalyocyanine
Arylamide yellow (CI11680) plus
toluidine red
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Table 11-3 (cont'd). Color Change After 12 Weeks of Exposure to a
Mixture of Photochemical Oxidants
Colorant*
Color Change (A£" units)f
Chemical Functionality or
Chemical Composition
Pararosaniline base (42500)*
Payne's Grey
Permanent Magenta
Permanent Rose
Prussian Blue
Prussian Green
Purple Lake
Reactive Blue 2 (61211)*
Rose Carthane (12467)
Rose Dore
Thioindigo Violet (73312)*
Winsor Yellow (11680)
25.6 ±4.7
1.0 ±0.1
l.liO.l
2.0 ±0.1
0.7 ±0.2
1.6 ±0.3
0.9 ±0.2
2.3 ±0.3
14.4 ± 1.1
0.8 ±0.2
2.0 ±0.2
1.9 ±1.2
0.5 ±0.2
Amino-substituted triphenylmethane
Alizarin lake plus prussian blue plus
lampblack plus ultramarine blue
Quinacridone
Quinacridone
Ferric ferrocyanide
Arylamide yellow plus prussian blue
Alizarin lake
Amino-substituted anthraquinone,
sulfonate salt
Arylamide (Pigment Red 10) plus
xanthene (Pigment Red 90)
Quinacridone plus Yellow 3
Chlorinated thioindigo
Arylamide yellow
* On watercolor paper unless otherwise indicated. Color Index (CI) names or CI numbers are given in
parentheses.
•f Mean ± one standard deviation for triplicate samples calculated from the parameters L*, a*, and b* measured
with the color analyzer.
{ On Whatman 41 paper.
Source: Grosjeanetal. (1993)
1 ozone-induce degradation. When ozone degradation was measured using scanning electron
2 microscopy, the oil house paint and latex coating samples showed erosion above that seen with
3 clean air but only at the highest exposure level. No effects were noted for the automotive paint.
4 The other painted surfaces were not evaluated.
5 Spence et al. (1975) studied the effect of air pollutants and relative humidity on oil based
6 house paint, acrylic latex house paint, acrylic coil coating, and vinyl coil coating under
7 laboratory conditions. Test panels were exposed in weathering chambers equipped with a xenon
8 are light for simulating sunlight to low and high levels of ozone (0.08 and 0.5 ppm), sulfur
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1 dioxide (0.03 and 0.5 ppm), and nitrogen dioxide (0.05 and 0.5 ppm) and relative humidity
2 (50 and 90%). Samples were exposed for a total of 1000 h. The exposure cycle consisted of
3 20 min of dew and 20 min of light. The effects of the exposure on the painted surfaces were
4 measured by weight loss and loss in film thickness. The acrylic coil coating had the lowest
5 erosion rate of the surface coatings tested. However, ozone was the only pollutant that had a
6 significant effect on the surface erosion. Sulfur dioxide and relative humidity were significant
7 factors in the erosion of oil base house paints and vinyl coil coating. The findings for acrylic
8 latex house paint were not reported.
9
10
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